Support Document for the Hazardous
Rule: Rlek Aeseseinent
for Human and
Volume I
Appendix B
Part 2 of 2
KtoZ
Prepared for
U.S. Envlronmenml Prolsctfon Agency
of Solid Wasts
Contr&el No.
August 1995
-------
APPENDIX B Kepone-1
Toxicological Profile for Selected Ecological Receptors
Kepone
Cas No.: 143-50-0
Summary: This profile on kepone summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are aJso
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
data presented in the technical support document for the Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including .aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
/
Mammals: No suitable subchronic or chronic studies were located in the literature for
mammalian wildlife in which dose-response data were reported. However, several chronic and
subchronic toxicity studies involving kepone have been conducted using laboratory rats and mice.
A reproductive study (Uphouse, 1986) was identified in which adult, female Fischer (F-344) rats
were injected intraperitoneally with 25, 50, or 75 mg/kg-diet of kepone (in cotton seed oil) on
the morning before mating or the morning after mating. Uphouse (1986) observed the number
of successful pregnancies, fertility, and litter size and recorded a NOAEL of 25 mg/kg-diet and
a LOAEL of 50 mg/kg-diet. Reproductive and chronic toxicity was observed in male and female
adult Sherman rats fed a dietary concentration of 25 ppm kepone for three months (Cannon and
Kimbrough, 1979). Cannon and Kimbrough (1979) observed a complete reproductive failure of
females and enlarged .livers in both sexes at 25 ppm. In a subchronic study, Chemoff and Rogers
August 1995
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APPENDIX B Kepone-2
(1976) reductions were observed in fetal weight, reduced degree of ossification, edema,
undescended testis, enlarged renal pelvis, and enlarged cerebral ventricles in the male rats
exposed in utero to kepone. They reported a NOAEL of 2 mg/kg-day and a LOAEL of 6 mg/kg-
day. In addition, Chernoff and Rogers (1976) observed increased fetal mortality and clubfoot in
the mouse corresponding to a NOAEL of 4 mg/kg-day and a LOAEL of 8 mg/kg-day. Larson
et al. (1979) fed 40 young male and 40 young female Wistar rats dietary concentrations of 1, 5,
10, 25, and 80 ppm of kepone for a period of 2 years. In this chronic toxicity study, Larson et
al. reported a LOAEL of 25 ppm for increased liver-to-body weight ratios, depressed growth,
elevated organ-to-body weight ratios for kidneys, spleen, heart, and testes, as well as degenerative
changes in liver cells, kidney lesions, and testicular atrophy. Based on the reference body weight
(kg) and the recommended value for food consumption (kg/day) for rats (U.S. EPA, 1988), the
LOAEL of 25 ppm was converted to 2 mg/kg-d and the 10 ppm NOAEL was converted to 0.76
mg/kg-day.
The NOAEL the Larson et al. (1979) study was chosen to derive the lexicological benchmark
because (1) exposures were administered via oral ingestion, (2) the study contained sufficient
dose-response information, and (3) the study had an experimentally derived NOAEL for a
developmental growth. The study by Chernoff and Rogers (1976) was not selected because (1)
the developmental endpoints of this study were considered less important to population
sustainability than the endpoints in the Larson et al. (1979) study and (2) the study was
considered subchronic. The study by Cannon and Kimbrough (1979) was not selected because
it lacked dose response information and was subchronic (3 months). The study by Uphouse
(1986) was not selected because of the uncertainity associated with extrapolating an ijection
exposure to a wildlife exposure. However, the aforementioned studies do illustrate the dose
ranges at which mammalian reproductive and development toxicity occurs from exposure to
kepone.
The selected NOAEL was then scaled for species representative of a freshwater ecosystem using
a cross-species scaling algorithm adapted from Opresko et al. (1994)
/ Benchmark^ = NOAEL. x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species', BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Larson et al.
(1979) study documented developmental effects from toxaphene exposure to female and male
rats, the mean body weight of both genders was used in the scaling algorithm to obtain the
lexicological benchmarks.
Data were available on reproductive and developmental effects, as well as growth or chronic
survival. In addition, the data set contained studies which were conducted over chronic and
subchronic durations and during sensitive life stages. All of the studies identified were conducted
August 1995
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APPENDIX B Kepone - 3
using laboratory rats or mice, and as such inter-species toxicity differences were not identifiable.
Thus, an inter-species uncertainty factor was not applied. There were no other toxicity values
in the kepone mammalian data set which were lower than the benchmark value. Therefore, based
on the data set for kepone, the benchmarks developed from the Larson et al. (1979) study were
categorized as adequate.
Birds: A subchronic study was identified in which male Japanese quail (Eroschenko, 1978) were
fed a dietary concentration of 200 ppm of kepone (suspended in acetone) for a period of 42 days.
Following this dietary concentration, Eroschenko (1978) noted the following structural changes
to the male reproductive organs: enlarged and atrophic testes and structural reparations in the
testes and ducts. In a reproductive study (Naber and Ware, 1965), hens were fed. dietary
concentrations of 75 or 150 ppm of,kepone for a 16-week period. Naber and Ware (1965) noted
a significant reduction in egg production and a reduction in the survival of chicks after hatching
at the 75 ppm concentration while lethality was noted at the 150 ppm concentration. A chronic
reproductive study was identified in which five week old Japanese quail were fed a diet
containing 10, 40, 80, and 160 ppm of kepone (suspended in sesame oil) for 250 days
(Eroschenko and Hackmann, 1981). Eroschenko and Hackmann (1981) observed ovulation, egg
production, egg laying, and egg quality and recorded a NOAEL of 80 ppm and a LOAEL of 160
ppm. A NOAEL of 8.74 mg/kg-day (80 ppm) was calculated based on the reference body weight
(kg) (Roseberry and Klimistra, 1971) and the value for food ingestion (kg/day) calculated from
the, following allometric equation Nagy (1987):
Food intake = 0.648 W 0-651, where W = body weight in grams
The NOAEL of 8.74 mg/kg-d reported by Eroschenko and Hackmann (1981) was used to
calculate the toxicological benchmark for birds because: (1) chr onic exposures were
administered via oral ingestion, (2) it focused on reproductive toxicity as a critical endpoint, and
(3) the study contained dose-response information. The study by Naber and Ware (1965) on
chickens lacked sufficient dose-response data. Likewise, the study by Eroschenko (1978) on
Japanese quails lacked dose-response information, in addition to being subchronic.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian species
representative of a freshwater ecosystem, the NOAEL of 8.74 mg/kg-day from Eroschenko and
Hackmann (1981) was scaled using the cross-species scaling method of Opresko et al. (1994).
Since the benchmark study documented reproductive effects from Kepone exposure to female
Japanese quail, female body weights for each representative species were used in the scaling
algorithm to obtain the toxicological benchmarks.
Data were available on the reproductive and developmental effects of kepone, as well as on
growth or survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations. Laboratory experiments of similar types were not conducted
on a range of avian species and as such, inter-species differences among wildlife species were
not identifiable. There were no other toxicity values in the avian data set which were lower than
the benchmark value. Based on the avian data set for kepone and the NOAEL from Eroschenko
August 1995
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APPENDIX B Kepone • 4
and Hackmann (1981), the benchmarks for avian species in the freshwater ecosystem were
categorized as adequate.
Fish and aquatic invertebrates: A review of the literature revealed that no AWQC existed for
kepone. Therefore, a Secondary Chronic Value (SCV) of 3.2E-4 mg/1 was calculated using the
Tier n methods described in Section 4.3.5. Since the benchmark for fish and aquatic
invertebrates was based on a SCV established using the Tier II methodology, it is categorized as
interim.
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwn). Aquatic
plant data was not identified for kepone and, therefore, no benchmark was developed.
Benthic Community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbon partition coefficient (K^) to determine a protective sediment concentration (Stephan,
1993). The EQp number is the chemical concentration that may be present in sediment while still
protecting the benthic community from the harmful effects of chemical exposure. Since neither
a FCV nor an AWQC exist for kepone, a Secondary Chronic Value (SCV) was calculated as
described in Section 4.3.5. The SCV reported for kepone was used to calculate an EQp number
of 9.67 mg kepone /kg organic carbon. Assuming a mass fraction of organic carbon for the
sediment (foc) of 0.05, the benchmark for the benthic community is 0.483 mg/kg. Since the EQp
number was based on a SCV established using the Tier II methodology, the sediment benchmark
is categorized as interim.
August 1995
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APPENDIX B
Kepone - 5
Table 1. Toxlcological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
ftepteaentallv*
Specie*
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
VaJuB« mo/Kg.
day
0.57 (a)
0.34 (a) -
4.06 (a)
5.00 (a)
4.57 (a)
5.44 (a)
6.07 (a)
12.12 (a)
5.68 (a)
9.12 (a)
Study
Specie*
rat
rat
Japanese
quail
Japanese
, quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Effect
dev
dev
rep
rep
rep .
rep
rep
rep
rep
rep
Study Value
mg/kg^tay
0.76
0.76
8.74
8.74
8.74
8.74
8.74
8.74
i
8.74
8.74
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
-
-
-
-
•
-
-
-
-
Original So we*
Larson et al., 1979
Larson el a)., 1979
Eroschenko and
Hackmann,. 1981
Erosohenko and
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Eroschenko and
Hackman. 1981
Eroschenko and
Hackmann, 1981
'Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Kepone - 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ftepr»*«matjv*
Sp«c3M
fish and aquatic
invertebrates
aquatic plants
benthic community
Benchmark
V*tu«*
oi^L
3.2 E-04 (i)
ID
0.483 (i)
mg/kg
sediment
Study Specie*
AWQC Species
AWQC Species
Description
scv
SCVx K.,.
Ofjfltoal $OBT5* :
AQUIRE, 1995
AQUIRE. 1995
IL
'Benchmark Category, a = adequate, p = provisional, i = interim; a '" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to kepone.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study,(Larson et al., 1979) was,used to derive the kepone toxicological benchmark for
mammalian species representing the terrestrial ecosystem. The NOAEL selected to be
representative of mammals in the generic terrestrial ecosystem was 0.79 mg/kg-day. This
value was then scaled for species representative of a terrestrial ecosystem using a cross-
species scaling algorithm adapted from Opresko et al. (1994). Since the Larson et al. (1979)
study documented developmental effects from toxaphene exposure to female and male rats,
the mean body weight of both genders was used in the scaling algorithm to obtain the
toxicological benchmarks. Based on the data set for kepone, the benchmarks developed from
the Larson et al. (1979) study were categorized as adequate.
Birds: No additional avian toxicity studies were identified for species representing the
terrestrial ecosystem. Although the minimum data set requirements of at least three test
species were not met, the NOAEL of 8.74 mg/kg-day established by Eroschenko and
Hackmann (1981) was used as the basis for the benchmark values. This NOAEL was then
scaled for species representative of a terrestrial ecosystem using a cross-species scaling
August 1995
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APPENDIX B Kepone-7
algorithm adapted from Opresko e.t al. (1994). Since the benchmark study documented
reproductive effects from kepone exposure to female Japanese quail, female body weights for
each representative species were used in the scaling algorithm to obtain the lexicological
benchmarks. Based on the avian data set for kepone and the NOAEL from the Eroschenko
and Hackmann (1981) study, the benchmarks for avian species in the generic terrestrial
ecosystem were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lerigth. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for kepone and, as a result, a benchmark
could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
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APPENDIX B
Kepone - 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
#*pr«**HtBliv«
SpaoiM
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
Etoftetinxtfc
Vaht**
mprtca-d
1.55 (a)
1.60 (a) .
1.35 (a)
0.55 (a)
0.39 (a)
0.37 (a)
0.19 (a)
5.47 (a)
9.61 (a)
8.74 (a)
10.58 (a)
8.80 (a)
ID
ID
Study
: speck*
rat
rat
rat
rat
rat
rat
rat
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
-
E««ct
rep
rep
rep
rep
rep
rep
rep
rep
. rep
rep
rep
rep
- .
Study
Valuo
raa/Xg-d
0.76
0.76
0.76
0.76
0.76
0.76
0.76
8.74
8.74
8.74
8.74
8.74
-
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
SF
-
•
•
•
•
QtfflJwiJ SOW!*
Larson et at.,
1979
Larson et at.,
1979
Larson et at.,
1979
Larson et al.,
1979
Larson et at.,
19791
Larson et ai.,
1979
Larson et at.,
1979
Eroschenko and
Hackmann, 1981
Eroschenko and .
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Eroschenko and
Hackmann, 1981
'Benchmark Category, a - adequate, p = provisional, i = interim; a "
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
indicates that the benchmark value was an order of
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
August 1995
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APPENDIX B Kepone-9
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in .section 5.3.2, the BAF/s for cohsituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for kepone. The
bioconcentration factor for fish was also estimated from the Thomann models (i.e., log Kow ~
dissolved BCF/) and multiplied by the dissolved fraction (/~d) as defined in Equation 6-21 to
determine the total bioconcentration factor (BCF/), The dissolved bioconcentration factor
(BCF/1) was converted to the BCF/ in order to estimate the acceptable lipid tissue
concentration (TC/) in fish consumed by piscivorous fish (see Equation 5-115). The BCF/
was required in Equation 5-115 because the surface water benchmark (i.e., FCV or SCV)
represents a total water concentration (C1). . Mathematically, conversion from BCF/1 to BCF/
was accomplished using the relationship delineated in the Interim Report on Data and
Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Wildlife (U.S.
EPA, 1993i): .
BCF/1 x fd = BCF/
As expected, converting the predicted BCF/1 of 31,623 L/kg LP to the BCF/ of 28,761 L/kg
LP was in good agreement (i.e., within a factor of about 0.5) with other predicted BCF/
values derived using regression equations (e.g., Veith and Kosian, 1983; Isnard and Lambert,
1989). The similarity in the BCF/1 and the BCF/, illustrates the trend in dissolved vs. total
water concentrations at lower log Kow values; as log Kow approaches 4.0, the dissolved
concentration is approximately equal to the total water concentration.
The bioaccumulation factor for terrestrial vertebrates, and the bioconcentration factors for
earthworms and invertebrates were estimated as described in Section 5.3.5.2.3. Briefly, the
extrapolation method is applied to hydrophobic organic chemicals assuming that the
partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs and
BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
August 1995
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APPENDIX B Kepone - 10
chemical lacking measured data (in this case dieldrin) is compared to the BBF for TCDD and
that ratio (i.e., kepone BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial
vertebrates, invertebrates, and earthworms, respectively. For hydrophobic organic
constituents, the bioconcentration factor for plants was estimated as described in Section 6.6.1
for above ground leafy vegetables and forage grasses. The BCF is based on rqute-to-leaf
translocation, direct deposition on leaves and grasses, and uptake into the plant through air
diffusion.
August 1995
-------
APPENDIX B
Kepone - 11
Table 4. Biological Uptake Properties
•colofllcal
receptot
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
tipid-ba*ed or
whole-body
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
value
35.849 (d)
35,710 (d)
28.761 (t)
33.475 (d)
35.541 (d)
71,181 (d)
0.00039
0.00037
0.003
0.097
•oure*
predicted value based on
Thomann, 1989, food chain
model
predicted value based on
Thomann, 1 989, food chain
model
predicted value based on
Thomann, 1989 and adjusted to
estimate total BCF
predicted value based on
Thomann et al.. 1992, food web
model •
predicted value based on
Thomann et al., 1992, food web
model
predicted value based on
Thomann et al., 1992, food web
model
estimated based on beef
biotransfer ratio with 2,3,7.8-
TCOO
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
estimated based on beef
biolransfer ratio with 2.3.7,8-
TCDD
U.S. EPA, 19929
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Kepqne - 12
References
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TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
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Buckler, D.R., A. Witt, F.L. Mayer, and J.N. Huckins., 1981. Acute and chronic effects of
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August 1995
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APPENDIX B Kepone-13
Johnson, W.W. and M.T. Finley. 1980. Handbook of Acute Toxicity of Chemicals to Fish
and Aquatic Invertebrates. Resource Publication 137, Fish and Wildlife Service, U.S.
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subchronic, and chronic toxicity of chlordecone. Toxicol. Appl. Pharmacol. 48:29-41.
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McFarland, L.Z., and P.B. Lacy. 1969. Physiological and endocrinological effects of the
insecticide Kepone on the Japanese quail. Toxicol. Appl. Pharmacol. 15:441-45.
Naber, E:C, and G.W. Ware. 1965. Effect of kepone and mirex on reproductive
performance in the laying hen. Poultry 'Sci. 44:875-880."
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
Opresko, D.M., B.E. Sample, and G.W. Suter II. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge
National Laboratory, Oak Ridge, TN.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Roberts, M.H. and R.E. Bendl. 1982. Acute toxicity of kepone to selected freshwater fishes.
Estuaries. 5(3): 158-164. As cited in AQUIRE (AQU&iic Toxicity /nformation /?£trieval
Database), Environmental Research Laboratory, Office of Research and Development,
U.S. Environmental Protection Agency, Duluth, MN
Roseberry, J.L., and W.D. Klimstra. 1971. Annual weight cycles in male and femal
bobwhite quail. Auk. 88:116-123.
Sanders, H.O., J. Huckins, B.T. Johnson, and D. Skaar. 1981. Biological Effects of Kepone
and Mirex in Freshwater Invertebrates. Archives of Environmental Contamination and
Toxicology, 10:531-539. As cited in AQUIRE (AOUaiic Toxicity/nformation flEtrieval
Database), Environmental Research Laboratory, Office of Research and Development,
U.S. Environmental Protection Agency, Duluth, MN
August 1995
-------
APPENDIX B Kepone - 14
Skaar, D.R., B.T. Johnson, J.R. Jones, and J.N. Huckins. 1981. Fate of kepone and mirex in
a model aquatice environment: sediment, fish, and diet Can. J. Fish. Aquat. Sci.,
38(8):931-938. As cited in AQUIRE C4g£/atic Toxicity /nformation tfEtrieval Database),
Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection 'Agency, Duluth, MN
Stephan, C.E. 1993. Derivations of proposed human health and wildlife bioac cumulation
factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter, G.W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic biota: 1994 Revision. ES/ER/TM-96/R1.
U.S. Department of Energy, Oak Ridge National Laboratory, Oak Ridge, TN.
U.S. EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for use in Risk Assessment. P338-179874.
Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1990. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, DC.- January. As cited in Pierson,
T.K., A.E. Crook, S.M. Beaulieu, P.N. Graham, N.B. Jones, A.M. Reynolds, and G.P.
Vegh, 1994, Development of Human Health Based Exit Criteria for the Hazardous Waste
Identification Project, Phase III Analysis.
U.S. EPA (Environmental Protection Agency). 1993. Technical Basis for Deriving Sediment
Quality Criteria for Nonionic Organic Contaminants for the Protection of Benthic
Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of Water,
Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1993i. Interim Report on Data and
. Methods for Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Life and
Associated Wildlife. EPA/600/R-93/055. Office of Research and Development,
Washington, DC.
Uphouse, L. 1986. Single injection with chlordecone reduces behavioral receptivity and
fertility of adult rats. Neurobeh. Toxicol. Tetratol. 8:121-126.
Vanveld, P.A. 1980. Uptake, Distribution, Metabolism, and Clearance of Kepone by Channel
Catfish (Ictalurus punctatus). M.A. Thesis, College of William and Mary, Williamburg,
VA. As cited in AQUIRE (AOUatic Toxicity_/nformation /?£trieval Database),
Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, Duluth, MN
^
August 1995
-------
Terrestrial Toxicity - Kepone
Cas No. 143-50-0
Name
kepone
,
kepone
kepone
kepone
kepone
kepone
kepone
kepone
Species
rat
rat
mouse
mouse
Wistar rats
Wistar rats
Sprague-
Oawley rats
adult female
Fisher rats
Endpolnt
dev
dev
dev
dev
dev
dev
^
rep
rep
Description
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
2
6
4
8
0.79
-
2
30
50
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
ppm
mq/kq-diet
Exposure
Route (oral,
8.C., I.V., l.p.,
Infection)
oral
oral
oral
oral
oral
•
oral
oral
•P
Exposure
Duratton/Tlmlnq
Days 7- 16 of
gestation
Days 7- 16 of
gestation
Days 7- 16 of
gestation
Days 7- 16 of
gestation
2 years
2 years .
90 days
once
Reference
Chernoff and Rogers,
1976
Chernoff and Rogers.
1976
Chernoff and Rogers,
1976
Chernoff and Rogers,
1976
Larson et al., 1979
Larson et al., 1979
Under el al., 1983
Uphouse, 1986
Comments
Fetotoxicity was not observed at
this dose level.
Administration of this dose level
resulted in general fetotoxicity.
i.e.. decreases in fetal weight
and the number of caudal
ossification centers.
Fetotoxicity was not observed at
this dose level.
Fetal toxic effects were noted at
dose levels which caused
considerable maternal toxicity
(as was the case with rats in the
same study)
No effects were observed at this
dose level.
Effects included the following:
minimal to severe testicular
atrophy.
The number of fertile males, the
number of litters, litter size, pup
viability, and fetal weight were
not affected at this dose level.
(30 ppm = 2.4 mg/kg-day)
Significant reduction in the
number of successful
pregnancies.
-------
APPENDIX B Kepone - 15
Will, M.E. and G.W.'Suter, 1994. lexicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
&<^Xtefoy&W^
August 1995
-------
Terrestrial Toxicity - Kepone
Cas No. 143-50-0
Chemical
Name
kepone
kepone
kepone
kepone
kepone
Species
rat
dog
rabbit
male
mallards
quail
Endpolnt
mort.
mort.
mort. i
mort.
mort.
Description
LD50
LD50
LD50
LD50
LD50
Value
95
250
65
167
237
Units
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg
mg/kg-body
wt.
Exposure
Route (oral,
B.C., I.V.. l.p.,
Infection)
oral
oral
oral
oral
oral
Exposure
Duration/Timing
NS
NS
NS
NS
NS
Reference
RTECS, 1994
RTECS, 1994
RTECS, 1994
Hudson, 1984
RTECS, 1994
Comments
NS = Not specified
-------
Terrestrial 1. jity - Kepone
Cas No. 143-50-0
Chemical
Name
kepone
kepone
kepone
kepone
kepone
kepone
kepone
kepone
Species
adult female
Fisher rats
male and
emale adult
Sherman
strain rats
Japanese
quail males ,
young or
adult quail
young or
adult quail
five-week old
Japanese
quail
five-week old
Japanese
quail
egg
production
type laying
hens
Endooint
rep
rep
rep, dvp,
mortality
rep, dvp
dev
rep
rep
rep
Description
NOAEL
AEL
AEL
AEL
AEL
NOAEL
LOAEL
LOAEL
Value
25
25
200
200
300-
80
160
75
Units
mg/kg-diet
ppm
ppm
ppm
ppm
ppm
ppm
ppm
Exposure
Route (oral,
S.C.. I.V.. l.p.t
Intectlon)
i.p.
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
once
3 -month feeding
period
Continuous
ingestion for 42
days
v
21 days
3 weeks
250 days
250 days
16 week period
Reference
Uphouse, 1986
.
Cannon el al., 1979
Eroschenko. 1978
Eroschenko and Wilson,
1975
McFartand and Lacy,
1969
Eroschenko and .
Hackmann, 1981
Eroschenko and
Hackmann, 1981
Naber and Ware, 1965
Comments
No reduction in fertility or liner
size.
'short-term exposure of dietary
chlordecone at the
concentrations tested does not
permanently affect reproduction
in rodents.'
'Gross and histological changes
in the quail testes depend not
only on the concentration but
also on the length of ingestion.'
'Produced uniformly enlarged,
edematous testes with tubular
distention and cellular debris'
'Produced both atrophic as well
as edematous testes.'
Reproductive effects were not
observed at this dose level.
Effects at this dose level
included the following:
increased quail mortality, a lag ir
egg production, and an effect on
the normal sequence of egg
laying.
There was a significant
reduction in egg production,
hatchability of eggs, and surviva
of chicks at this dose level. (15C
ppm =lethalitv)
-------
Freshwater Biological Uptake Measures • Kepone
Cas No. 143-50-0
Chemical
Name
kepone
kepone
kepone
Spec lea
channel
catfish
channel
catfish
blueaill
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
Value
1,163
3
10.606
Measured
or
predicted
(m,p)
m
m
m
Unite
NS
NS
NS
Reference
Roberts et al., 1982 as cited In
AQUIRE, 1995
Vanveld, 1980 as cited in
AQUIRE, 1995
Skaar et al., 1981 as cited in
AQUIRE. 1995
Comments
Juvenile; 4 day test.
Day 6.5 of gestation; 90
day test.
Days 0.5 -15 of gestation;
7 - 28 day test.
NS = Not specified
-------
Freshwater\ acity - Kepone
CAS No. 143-50-0
Chemical
Name
kepone
kepone
kepone
kepone
Species
Daphnia
magna
channel
catfish
bluegill
fathead
minnow
Effect (dvp,
rep, emb,
fet, behv,
. chron,
acute)
immob
mort.
mort.
mort.
Description
EC50
LC50
LC50
LC50
Value
260
422-512
(467.74)
30-66
(49.69)
340-420
(375.8)
Units '
ug/L
ug/L
ug/L
uoyu
Test type
(static/ flow
throuqh)
NS
NS
NS
NS
Exposure
Duration/
Timing
48 hour
96 hour
96 hour
4 days
Reference
Sanders et al., 1981 as cited
InAQUIRE, 1995
Roberts et al., 1982 as cited
inAQUIRE, 1995
Roberts et al., 1982 as cited
inAQUIRE, 1995
Buckler et al., 1981 as cited
inAQUIRE, 1995
Comments
NS = Not specified
-------
Terrestrial Biological i jke Measures - Kepone
Cas No. 143-50-0
Chemical
Name
Kepone
Species
plants
B-factor
(BCF, BAF,
BMP)
BCF
Value
69
Measured
or
predicted
(m.p)
p
Units
(ug/g WW plant)/(ug/mL
soil water)
Reference
U.S. EPA, 1990e
Comments
-------
APPENDIX B Lead - 1
lexicological Profile for Selected Ecological Receptors
Lead
Cas No.: 7439-92-1
Summary: This profile on lead summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the
freshwater ecosystem were calculated for organic constituents with log Kow between 4 and
6.5. For the terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire lexicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most
current information and may differ from data presented in the technical support document for
the Hazardous Waste Indentification Rule (HWIR): Risk Assessment for Human and
Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concemrations (C ) for ihe generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Numerous daia were identified concerning the effects of lead toxiciry in mammals.
In an experiment lasting 20-30 days, rats were administered lead in oral doses of 0.05, 0.005
and 0.0015 mg/kg-day (Krasovskii el al., 1979). Impairmeni of ihe functional capacity of the
male rat's spermatozoa was observed in rals receiving ihe maximum dose of 0.05 mg/kg-day.
The gonadoioxic effecls al 0.05 mg/kg-day resuhed in an inferred NOAEL of 0.005 mg/kg-
day. In anoihef experimeni in ihe study, male and female rats were given ihe same doses of
lead mentioned above for 6-12 months. Neurological deficits, including disruption of
conditional responses and motor activity, were observed al 0.05 and 0.005 mg/kg-day. In
another investigation, dogs given a single dietary dose of 0.32 mg/kg-day for an unspecified
period of time exhibited clinical signs of chronic lead toxicity (Demayo et al., 1982). Also,
Hilderbrand el al. (1973) ireated male and female rals lo oral doses of 5 and 100 ug/day of
August 1995
-------
APPENDIX B Lead - 2
lead for 30 days. In this study gonadotoxic effects in both the male and female rats were
observed at the 100 ug/day dose resulting in a NOAEL of 5 ug/day. This effects level
corresponds to a daily dose of 0.022 mg/kg-day. To obtain the NOAEL as a daily dose, the
reported dose was divided by the geometric mean (0.235kg) of the male and female rat's
reported body weights.
The NOAEL for gonadotoxic effects from the Krasovskii et ah, (1979) study was chosen to
derive the lexicological benchmark because (1) chronic exposures were administered via oral
ingestion, (2) it focused on irregularities in the male rat's reproductive system as a critical
endpoint, (3) the study contained dose response information, and (4) the study reported the
lowest toxicity value for a critical endpoint. The study by Hilderbrand et al., (1973) was not
selected for the derivation of a benchmark because it did not report the lowest toxicity value
for a critical endpoint. The Demayo et al., (1982) study was not chosen because of the
absence of sufficient dose-response information and lack of critical endpoints.
The study value from Krasovskii et al., (1979) was scaled for species representative of a.
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
(bw V4
Benchmarkw = NOAEL, x _ L
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Krasovskii et al. (1979) study documented reproductive effects from lead exposure to male
rats, the mean male body weight of representative species was used in the scaling algorithm
to obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of lead, as well as growth
or chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. The data set contained a
study value for neurological endpoints (Krasovskii et al., 1979) that was approximately an
order of magnitude lower than the benchmark value. Based on the data set for lead the
benchmarks developed from the Krasovskii et al. (1979) study were categorized, adequate,
with a "*" to indicate that some adverse effects have been observed at the benchmark level.
Birds: There were several studies that investigated the effects of lead toxicity on birds.
Growth rate suppression occurred in chickens exposed to 1850 ppm of dietary lead for 4
weeks (Franson and Custer, 1982). This level corresponds to a daily dose of 196 mg Pb/kg-
day based on the geometric mean body weight of 0.109 kg for the control birds in the study
and the derived food consumption rate of 0.0116 kg/day (U.S. EPA, 1988). American
kestrels exposed to doses of 10 and 50 ppm for 6 months exhibited no impairment of
August 1995
-------
APPENDIX B Lead - 3
survival, egg laying, fertility, or egg thickness (Pattee, 1984). The 50 ppm dose was
converted to a daily dose of 6.3 mg/kg-day based on the kestrel body weight of 0.119 kg
(U.S. EPA, 1993g) and the derived food intake rate of 0.015 kg/day (Nagy, 1987).
In another study, Hoffman et al., (1985) examined the growth of one-day old American
kestrel nestlings exposed orally to 25, 125 and 625 mg/kg-day of dietary lead. The authors
reported a NOAEL of 25 mg/kg-day and a LOAEL of 125 mg/kg-day. In a series of
experiments, Edens and Garlich (1983) monitored the egg production of chickens and
Japanese quail. The lowest effects level reported in their study resulted when newly hatched
Japanese quail were exposed to 1, 10, and 100 ppm of lead in their diets for 5 weeks. This
experiment resulted in a reported LOAEL of 1 mg/kg. This corresponds to a daily dose of
0.21 mg/kg-day based on a body weight value of 0.150 kg and a food intake value of 0.031
kg/day, both obtained from the study.
The LOAEL reported by Edens and Garlich (1983) for Japanese quail was selected to derive
the avian benchmark value for the freshwater ecosystem. This study was chosen because (1)
it focused on reproductive toxicity as the primary endpoint, (2) the study contained dose-
response information.and (3) the study contained the lowest toxicity value for a critical
endpoint. The other studies mentioned above were not selected, either because they did not
focus on a reproductive endpoint or they lacked sufficient dose-response information.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian
species representative of a freshwater ecosystem, the LOAEL of 0.21 mg/kg-day from the
Edens and Garlich (1983) study was divided by 10 to provide for a LOAEL to NOAEL safety
factor, and scaled using the cross-species scaling method of Opresko et al. (1994). Since the
Edens and Garlich (1983) study documented reproductive effects from lead on female
Japanese quail, female body weights for each representative species were used in the scaling
algorithm to obtain the lexicological benchmarks.
Data were available on reproductive and developmental effects as well as on growth or
survival. In addition, the data set contained studies that were conducted over chronic and
subchronic durations as well as during a sensitive life stage. There were no other values in
the data set below the benchmark value by at least an order of magnitude. Laboratory
experiments of similar types were not conducted on a range of avian species and, as such,
inter-species differences among wildlife species were not identifiable. Based on the avian
data set for lead, the benchmarks developed from the LOAEL in the Edens and Garlich
(1983) study were categorized as provisional.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) for lead of 3.2E-03 mg/L
was selected as the benchmark protective of fish and aquatic invertebrates (U.S. EPA, 1985).
The FCV for lead is based on the equation e^273^8"11165*)!-4-705). It is a hardness dependent
criterion that is normalized to 100 mg/L. Since the FCV was derived in the AWQC
document, the benchmark was categorized as adequate.
August 1995
-------
APPENDIX B
Lead -4
Aquatic plants: The benchmarks for aquatic plants were either (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants, (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). For lead the
benchmark value was determined to be 5.00E-K)2 mg/L based on the growth inhibition of
Chlorella vulgaris, Scenedesmus quadricauda and Selenastrum capricornutum. As described
in Section 4.3.6, all benchmarks for aquatic plants were designated as interim.
Benthic community: The lead benchmark protective of benthic organisms is pending a U.S.
EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
Raprewmlrfv*
OpiQtM
mink
river otter
bald Mete
osprey
groat blue heron
mallard
lesser scaup
spotted sandpiper
herring gul
kingfisher
OeBjClimUfl
V»to»*«a*B-
«t*y
0.003 (a*)
0.002 (a*)
0.006 (p)
0.007 (p)
0.007 (p)
0.008 (p)
0.009 (p)
0.019 (p)
0.009 (p)
0.014 (p)
Stody
dfMMBlMt
rat
rat
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
Japanese
quail
fHMt
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
awdyvafc*
«9*HV
0.005
0.005
0.21
0.21
0.21
0.21
0.21
0.21
0.21
0.21
.... -.-. ^
OMMtpilon
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
m
10
10
10
10
10
10
10
10
CMtofcrtfcwn*
s f -f --s<.
Krasovskiietal.,
1979
Krasovskii et a)..
1979
Edens and Gariich,
1983
Edens and Gariich,
1983
Edens and Gartich,
1963
Edens and Gartich,
1983
Edens and Gartich,
1983
Edens and Gariich,
1963
Edens and Gariich,
1963
Edens and Gartich,
1983
'Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Lead-5
Table 2, Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
R*pr***m0dv*
SttAfii^*
^rypwc^^v
fish and aquatic
invertebrates
aquatic plants
banlhic community
DeiiChflllilL
Vila**
mgfc
3.2E-03 (a)
5.0E+02 (i)
under review
Study
8|M<*M
aquatic
organisms
aquatic
plants
-
||^H^^V*pqn^V
FCV
CV
-
. Qrtgta«tS«««*
U.S. EPA. 1985
Suter and Mabrey,
1994
-
II.
•Benchmark Category, a = adequate, p = provisional, i = interim; a "' indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to lead. Because
of the lack of additional mammalian toxicity studies, the same surrogate-species study
(Krasovskii et al., 1979) was used to derive the lead lexicological benchmark for mammalian
species representing the terrestrial ecosystem. The study NOAEL of 0.005 mg/kg-day from
the Krasovskii et al. (1979) study was scaled for species in the terrestrial ecosystem using a
cross-species scaling algorithm adapted from Opresko et al. (1994). Since the Krasovskii et
al. (1979) study documented reproductive effects from lead exposure to male rats, the male
body weight of each representative species was used in the scaling algorithm to obtain the
lexicological benchmarks.
Based on the data set for lead the benchmarks developed from the Krasovskii et al. (1973)
study were categorized as adequate, with a "*" to indicate that some adverse effects have
been observed at the benchmark level.
Birds: As in the freshwater ecosystem, the study by Edens and Garlich (1983) was used to
calculate the benchmarks for birds in the generic terrestrial ecosystem. The study value of
1.0 mg/kg (0.21 mg/kg-day) from the Edens and Garlich (1983) study was divided by 10 to
provide a LOAEL-to-NOAEL safety factor. The LOAEL/10 was then scaled for the
representative species using the cross-species scaling algorithm adapted from Opresko et al.
August 1995
-------
APPENDIX B Lead-6
(1994). Since the Edens and Garlich (1983) study documented reproductive effects of lead
on female Japanese quail, female body weights for each representative species were used in
the scaling algorithm to obtain the lexicological benchmarks. Because the benchmarks were
derived from a LOAEL/10, they were categorized as provisional.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the LOEC values and then approximating the 10" percentile.
If there were 10 or fewer values, the 10* percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation. The
selected benchmark for phytotoxic effects of lead in soils is 50 mg/kg (Will and Suter, 1994).
Since the study value selected is the 10th percentile of more than 10 LOEC values, the
terrestrial benchmark for lead is categorized as provisional.
Soil Community: For the soil community, the lexicological benchmarks were established
based on methods developed by the Dutch National Institute of Public Health and
Environmental Protection (RIVM). In brief, the RIVM approach estimates a concentration at
which the no observed effect concentration (NOEQ for 95 percent of the species within the
community is not exceeded. A minimum data set was established in which key structural and
functional components of the soil community (e.g., microfauna, mesofauna, and macrofauna)
were represented. Measurement endpoints included reproductive effects as well as measures
of mortality, growth and survival. The derived lead benchmark deemed protective of the soil
community is 0.2534 mg/kg. Since the lead data set contains NOECs and/or LOECs for at
least five of the representative species outlined in the minimum data set, the soil community
benchmark is categorized as provisional.
August 1995
-------
APPENDIX B
Lead-7
Table 3. ToxicologicaJ Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Bepr*««nUHfr»
80*ci*0
doer mouse
short-tailed
shrew
meadow vote
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tatted hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
Value-
«*%•**
0.009 (a*)
0.010 (a*)
0.008 (a*)
0.003(a*)
0.002 (a')
0.002 (a*)
0.001 (a*)
0.006 (p)
0.014 (p)
0.013 (p)
0.016 (p)
0.013 (p)
50 (p) mg/kg
0.2534 (p)
mg/fcg
8Mdy
fitt^ifatt
•^p^^*^^^
rat
rat
rat
rat
rat
rat
rat
Japanese
quail
Japanese
puaM
Japanese
quail
Japanese
quail
Japanese
quail
terrestrial
plants
soil
invertebrat
es
£»•*
rep
rep
rep
rep.
rep
rep
rep
rep
rep
rep
rep
rep
growth/
yield
chronic
*0*
Value
•W
**
0.005
0.005
0.005
0.006
0.005
0.005
0.005
0.21
0.21
0.21
0.21
0.21
50 mg/kg
0.2534
mgAg
0wwdwton
"• s x
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
10thpercentile
LOEC
NOEC
: Wf
-
•
-
-
-
-
10
10
10
10
10
•
.- A^fiy^R^9 w^W^^^P
fff \
Krasovskji et al.,
1979
Krasovskii et al.,
1979
Krasovskii et al.,
1979
Krasovskii et al.,
1979
. Krasovskii et al.,
1979
Krasovskii et al.,
1979
Krasovskii et al.,
1979
Edens and Gartich,
1963
Edens and GarVch,
1963
Edens andGartch,
1963
Edens and Gariich,
1983
Edens and Gariich,
1983
Will and Suter.
1994
AJdenbergand
Slob, 1993
'Benchmark Category, a - adequate, p = provisional, i = interim; a "' indicates (hat the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Lead - 8
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. For
metals, BCFs are whole-body bioconcentration factors and refer to total surface water
concentrations (versus freely dissolved concentrations). Consequently all calculations of
acceptable tissue concentrations (TC) represent whole-body concentrations. The brief
discussion following Table 4 describes the rationale for selecting the biological uptake factors
and provides the context for interpreting the biological uptake values.
The whole-body BCF for lead was the geometric mean of 2 measured values; 42, supplied
by Stephan et al. (1993) and 46? supplied by USFWS (1988). BCF values for muscle were
not included because ecological receptors are likely to eat the whole fish, or in least, will not
necessarily distinguish between the fillet and other parts of the fish. Data on bioconcentration
in aquatic invertebrates are under review. Studies in bioconcentration/bioaccumulation in
terrestrial vertebrates and invertebrates have been identified and are currently being reviewed?
?? For metals, empirical data were used to derive the BCF for aboveground forage grasses
and leafy vegetables. In particular the uptake-response slope for forage grasses and was used
as the BCF for plants in the terrestrial ecosystem since most of the representative plant-eating
species feed on wild grasses
August 1995
-------
APPENDIX B
Lead - 9
Table 4. Biological Uptake Properties
Meto^ori
: flMMptVT
fish
littoraJ
trophic level 2
invertebrates
terrestrial • '
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF,lAP,«r
BSAF
BCF
-
BAF
BAF
BAF
BAF
KpMbwdgf
M^hfit^ hrt Ar
WlKHPvMwy'
whole
whol«-body
whole-body
whole-body
whole- plant
Mto.
44 (t)
ID
2.7E:01
3.2E-02
1.9E-01
2.4E-01
•»
-------
APPENDIX B Lead - 10
References
AQUIRE (AQUatic Toxicity /nformation flftrieval Database). 1995. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental
Protection Agency, Duluth, MN. July.
Aldenberg, T. and W. Slob. 1993. Confidence limits for hazardous concentrations based on
logistically distributed NOEC toxicity data. Ecotoxicology and Environmental Safety.
25:48-63.
Carson, T. L., G. A. Van Gelder, G. C. Karas, W. B. Buck. 1974. Slowed learning in
lambs prenatally exposed to lead. Archives of Environmental Health. 29:154-156.
i
Clark, D. R., Jr. 1979. Lead Concentrations: Bats. vs. terrestrial small mammals collected
near a major highway.. Environmental Science and Technology 13:338-341. As cited in
U.S. Department of the Interior, Fish and Wildlife Service, 1988, Lead Hazards to Fish,
Wildlife, and Invertebrates: A Synoptic Review, Biological Report 85(1.14).
Coughlan, D. J., S. P. Gloss, and J. Kubota. 1986. Acute and sub-chronic toxicity of lead to
the early life stages of smallmouth bass (Micropterus dolomieui). Water, Air, and Soil
Pollution 28:265-275.
Davies, P. H., J. G. Goettl, J. R. Sinley, Jr., and N. F. Smith. 1976. Acute and chronic
toxicity of lead to rainbow trout, Salmo gairdneri, in hard and soft water. Water Res.
10:199-206.
Davies, B. E., 1983. Heavy Metal Contamination from Base Metal Mining and Smelting:
Implications for Man and His Environment. Applied Environmental Geochemistry, ISBN
0-12-690640-8. As cited in RTI, 1991, Comparison of Risk Assessment Methodologies for
Selected Metals in Sewage Sludge, Research Triangle Park, NC.
Demayo, A., M. C., K. W. Taylor, and P. V. Hodson. 1982. Toxic effects of lead and lead
compounds on human health, aquatic life, wildlife plants, and livestock. CRC Critical
Reviews in Environmental Controls 12:257-305.
Edens, F.W., and J. D. Garlich,. 1983. Lead-induced egg production decrease in Leghorn
and Japanese quail hens. Poultry Science. 62:1557-1763.
57 FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg3/4/day.
August 1995
-------
APPENDIX B Lead - 11
Franson, J. C, and T: W. Custer. 1982. Toxicity of dietary lead in young cockerels.
Veterinary and Human Toxicology 24:421-423. As cited in U.S. Department of the
Interior, Fish and Wildlife Service, 1988, Lead Hazards to Fish, Wildlife, and
Invertebrates: A Synoptic Review, Biological Report 85(1.14).
Franson, J. C., L. Sileo, O. H. Pattee, and J. F. Moore. 1983. Effects of chronic dietary lead
in American kestrels (Falco sparverius). J. Wildl. Dis. 19:110-113.
Hale, J. G. 1977. Toxicity of metal mining wastes. Bulletin of Environmental
Contamination and Toxicology. 17:66-73.
Hartenstein, R., E. F. Neuhauser, E. F., and J. Collier. 1980. Accumulation of heavy metals
in the earthworm Eisenia foetida. Journal of Environmental Quality 9:23-26. As cited in
Environmental Health Criteria 85. Lead — Environmental Aspects, World Health
Organization, Geneva, 1989. •
Hilderbrand, D. C., R. Der, W. T. Griffin, et al. 1973. Effect of lead, acetate on
reproduction. Am J. Obstet Gynecol 115:1058-1065. As cited in Toxicological Profile
for Lead, Agency for Toxic Substances and Disease Registry, U.S. Public Health Service,
1993.
Hoffman, D. J., J. C. Franson, O. H. Pattee, C. N. Bunck, and A. Anderson. 1985. Survival,
growth and accumulation of ingested lead in nesting American kestrels (Falco sparverius).
Archives of Environmental Contamination and Toxicology 14:89-94.
Holcombe, G. W., D. A. Benoit, E. N. Leonard, and J. M. McKim. 1976. Long-term effects
of lead exposure on three generations of brook trout (Salvelinus fontinalis). J. Fish Res.
Board Can. 33:1731-4L As cited in Spry, D. J. and J. G. Wiener, 1991, Metal
bioavailability and toxicity to fish in low-alkalinity lakes: a critical review,
Environmental Pollution 71:243-304.
Kimmel, C. A., L. D: Grant, C S. Sloan, and B. C. Gladen. 1980. Chronic low-level lead
toxicity in the rat. I. Maternal toxicity and perinatal effects. Toxicology and Applied
Pharmacology 56:28-41. As cited in U.S. Department of the Interior, Fish and Wildlife
Service. 1988. Lead Hazards to Fish, Wildlife, and Invertebrates: A Synoptic Review,
Biological Report 85(1.14).
Krasovskii, G. N., L. Y. Vasukovish, and O. G. Chariev. 1979. Experimental study of
biological effects of lead and aluminum following oral administration. Environ Health
Perspect 30:47-51.
Ma, W. 1989. Effect of soil pollution with metallic lead pellets on lead bioaccumulation and
organ/body weight alterations in small mammals. Archives of Environmental
Contamination and Toxicology. 18:617-622.
August 1995
-------
APPENDIX B Lead - 12
Merlini, M. and G. Pozzi. 1977. Lead and freshwater fishes part I. Lead accumulation and
water pH. Environmental Pollution 12:167-172.
Muramoto, S. 1980. Effect of complexans (EDTA, NTA and DTPA) on the exposure of
high concentrations of cadmium, copper, zinc and lead. Bulletin of Environmental
Contamination and Toxicology 25:941-946.
Nagy, K. A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
EcolMono 57:111-128.
Naqvi, S. M., R. D. Howell, and M. Sholas. 1993. Cadmium and lead residues in field-
collected red swamp crayfish (Procambarus Clarkii) and uptake of alligator weed,
Alternanthera philoxiroides. Journal of Environmental Science and Health. B28(4):473-
485.
NRCC. 1973. Lead in the Canadian Environment. Natl. Res. Coun. Canada Publ. BY73-7
(ES). 116 pp.-Avail, from Publications, NRCC/CNRC, Ottawa, Canada K1A OR6.
NRCC. 1978. Effects of lead in the environment - 1978: quantitative aspects. Natl. Res.
Coun. Canada Publ. NNRC/CNRC
Nriagu, J. O. (ed.). 1978. The Biogeochemistry of Lead in the Environment. Part B.
Biological Effects. Elsevier/North Holland Biomedical Press, Amsterdam. 397 pp. As
cited in U.S. Department of the Interior, Fish and Wildlife Service, 1988, Lead Hazards to
Fish, Wildlife, and Invertebrates: A Synoptic Review, Biological Report 85(1.14).
Opresko, D. M., B. E. Sample, and G. W. Suter. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1.
Pattee, O. H. 1984. Eggshell thickness and reproduction in American kestrels exposed to
chronic dietary lead. Archives of Environmental Contamination and Toxicology 13:29-34.
Reiser M. H., S. A. Temple. 1983. Effects of chronic lead ingestion on birds of prey. In:
Recent Advances in the Study of Raptor Diseases. Proceedings of the International
Symposium on Diseases of Birds of Prey. July 1-3, 1980. 21-25. Chiron Publications.
Reiter, L. W., G. E. Anderson, J. W. Laskey, et al. 1975. Development and behavioral
changes in the rat during chronic exposure to lead. Environ. Health Perspect 12:119-123.
As cited in Environmental Health Criteria 85. Lead — Environmental Aspects, World
Health Organization, Geneva, 1989.
Roth, M. 1993. Investigations on lead in the soil invertebrates of a forest ecosystem.
Pedobiologia. 37:270-279.
August 1995
-------
APPENDIX B Lead. 13
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Scheuhammer, A. M. 1987. The Chronic Toxicity of Aluminam, Cadmium, Mercury, and
Lead in Birds: A Review. Environmental Pollution. 46:263-295.
Sharma, R. M. and W. B. Buck. 1976. Effects of chronic lead exposure to pregnant sheep
and their progeny. Veterinary Toxicology. 18:186-188.
Singhal R. L. and J. A. Thomas. 1980. Lead Toxicity. Effects of Lead on Mammalian
Reproduction. Urban and Schwartzenburg, Inc., Baltimore-Munich.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN, PB93-154672.
Suter n, G.W., M.A. Futrell, and G.A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
Suter n, G.W., and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-
96/R1.
Tachon, P., A. Laschi, J. P. Briffaux, and G. Brian. 1983. Lead poisoning in monkeys
during pregnancy and lactation. The Science of the Total Environment 30:221-229.
Taylor, D. H., C. W. Steele, and S. Strickier-Shaw. 1990. Responses of green frog (Rana
clamitans) tadpoles to lead-polluted water. Environmental Toxicology Chemistry 9:87-93.
U.S. Department of the Interior, Fish and Wildlife Service. 1988. Lead Hazards to Fish,
Wildlife, and Invertebrates: A Synoptic Review. Biological Report 85(1.14).
U.S. EPA (Environmental Protection Agency). 1980. Ambient Water Quality Criteria for
Lead. U.S. Environmental Protection Agency Rep. 440/5-80-057. 151 pp. Avail, from
NTIS, 5285 Port Royal Road, Springfield, VA 22161. As cited in U.S. Department of
the Interior, Fish and Wildlife Service, 1988, Lead Hazards to Fish, Wildlife, and
Invertebrates: A Synoptic Review, Biological Report 85(1.14).
U.S. EPA (Environmental Protection Agency). 1985. Ambient Water Quality Criteria for
Lead—1984. U.S. Environmental Protection Agency Rep. 440/5-84-027. 81 pp. Avail.
from NTIS, 5285 Port Royal Road, Springfield, VA 22161.
August 1995
-------
APPENDIX B Lead - 14
U.S. EPA (Environmental Protection Agency). 1986. Health Effects Assessment for Lead.
U.S. Environmental Protection Agency Rep. 540/1-86-055. 43 pp. Avail, from NTIS,
5285 Port Royal Road, Springfield, VA 22161.
U.S. EPA (Environmental Protection Agency). 19881. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. EPA P338-179874. U.S.
EPA, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1990. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, DC. January.
U.S. EPA (Environmental Protection Agency). 1993g. Wildlife Exposure Factors Handbook:
Volumes I and II. EPA/600/R-93/187a,b. Office of Research and Development,
Washington, DC.
U.S. EPA; (Environmental Protection Agency). 1994. Integrated Risk Information System.
September.
Vighi, M. 1980. Lead Uptake and Release in an Experimental Trophic Chain.
Ecotoxicology and Environmental Safety. 5:177-193.
Wide, M. 1985. Lead exposure on critical days of fetal life affects fertility in the female
mouse. Teratology 32:375-380. As cited in U.S. Department of the Interior, Fish and
Wildlife Service, 1988, Lead Hazards to Fish, Wildlife, and Invertebrates: A Synoptic
Review, Biological Report 85(1.14).
Will, M. E., and G. W. Suter, II. 1994. Toxicological benchmarks for Screenig Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for the U. S. Department of Energy.
Wong, P. T. S., B. A. Silverberg, Y. K. Chau, and P. V. Hodson. 1978. Lead and the
aquatic biota, pp. 279-342. In J. O. Nriagu (ed.). The Biogeochemistry of Lead in the
Environment. Part B. Biological Effects. Elsevier/North Holland Biomedical Press,
Amsterdam.
Wong, P. T. S., Y. K. Chau, O. Kramar, and G. A. Bengert. 1981. Accumulation and
depuration of tetramethyllead by rainbow trout. Water Resources 15:621-625. As cited
in U.S. Department of the Interior, Fish and Wildlife Service, 1988, Lead Hazards to
Fish, Wildlife, and Invertebrates: A Synoptic Review, Biological Report 85(1.14).
Zmudzki, J. G. R. Bratton, C. Womac, and L. Rowe. 1983. Lead poisoning in cattle:
reassessment of the minimum toxic oral dose. Bulletin of Environmental Contamination
and Toxicology 30:435-441.
August 1995
-------
Freshwater Biologica .take Measures - Lead
Gas No. 7439-92-1
Chemical Name
lead (lead nitrate)
NS = Not specified
Species
pumpkinseed
B-factor
(BCF. BAF,
BMP)
BCF
Value
486
Measured
or
Predicted
(m,p)
m
Units
NS
Reference
Merlini and Pozzi,
1977
Comments
Exposed for 8 days to 40 ug/L;
whole body BCF based on a
radioactive tracer.
-------
Terrestrial Biological Uptake Measures - Lead
CAS No. 7439-92-1
Chemical Name
lead
lead (lead
acetate)
lead
lead
lead
lead
lead (metallic
lead)
lead (metallic
lead)
lead (triethyl
lead)
lead (trimethyl
lead)
Species
plant
earthworm
earthworm
earthworm
earthworm
insects
American
kestrel
nestling
American
kestrel
nestling
starling
starling
NS = Not specified
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
Measured
or
Predicted
(m.p)
0.02
0.07
27
067
001
500.00
0.084
005
0.65
1.9
m
NS
m
m
m
m
NS
NS
NS
NS
i
Units
mg/kg DW
of
plant)/(mg/kg
soil)
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
'
_
FWS, 1988
'
Hartenstein et al , 1980
Oayies, 1983
Davies, 1983
Davies, 1983
U.S. EPA, 1985
Hoffman et al , 1985 as cited in
WHO, 1989
Hoffman et al., 1985 as cited in
WHO. 1989
Osborn et al., 1983 as cited in
WHO, 1989
Osborn et al., 1983 as cited in
WHO, 1989
-
Comments
Exposure through sewage for 35 days
to 2500 mg/kg ; whole body BCF
Data taken from Cwmystwyth, Ireland.
Data taken from Borth, Ireland.
Data taken from Dolgellau. Ireland.
Oral exposure for 10 days to 25 mg/kg-
day; kidney BCF.
Oral exposure for 10 days to 25 mg/kg-
day; liver BCF.
Oral exposure for 1 1 days to 2.85
mg/kg-day: kidney, wet weight BCF.
Oral exposure for 1 1 days to 2.85
mg/kg-day; kidney, wet weight BCF
|
-------
Freshwate. xicity - Lead
CAS No. 7439-92-1
Chemical Name
lead
lead
lead
lead
lead
NA = Not applicable
NS = Not specified
Species
fish
smallmouth
bass
Daphnia
Magna
Daphnia pulex
rainbow trout
Type of Effect
chron
acute
acute
acute
acute
Description
EC20
LC50
LC50
LC50
TL50
Value
22
2200-
29,000
(5.623)
4400
5100
8
Units
ug/L
ug/L
ug/L
ug/L
mg/L
Test Type
(static/flow
through)
NA
static
static
static
flowthrough
Exposure
Duration/
Timing
NS
4 days
4 days
4 days
4 days
Reference
Suteretal., 1992
Coughlan et al., 1986
as cited in AQUIRE,
1994
Mount and Norberg,
1984 as cited in
Aquire,1995.
Mount and Norberg,
1 984 as cited in Aquire,
1995
Hale, 1977
Comments
•
Fish were 2 months old.
-------
Freshwater Biological Uptake Measures - Lead
Cas No. 7439-92-1
Chemical Name
lead
lead
lead
lead
lead
lead
lead
lead
lead
lead (lead nitrate)
lead (lead nitrate)
Species .
bluegill
red swamp
crayfish
rainbow trout
rainbow trout
rainbow trout
brook trout
brook trout
brook trout
brook trout
carp
carp
B-factor
(BCF, BAF.
BMP)
BAF
_BAF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
45.70
17°.
72600
17,300
12.540
571.00
1.806.00
42000
1,504.00
4.200.00
304.00
Measured
or
Predicted
(mip)
P
m
NS
NS
NS
NS
NS
NS
NS
"1
m
Units
NS
_NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1993
Naqvietal ,1993
Wongetal., 1981
Wong et al.. 1981
Wongetal,, 1981
Wongetal.. 1978
as cited in FWS,
1988
Wongetal., 1978
as cited in FWS,
1988
Wongetal., 1978
as cited in FWS,
1988 ____
Wongetal., 1978
as cited in FWS,
1988
Muramoto, 1980 as
cited in WHO,
1989
Muramoto, 1980 as
cited in WHO,
1989
Comments
Whole body BAF.
BCF for whole trout.
BCF for intestinal lipids at day 10
of exposure.
BCF for intestinal lipids at day 14
of exposure
Study conducted over 3
generations; BCF for liver of first
generation.
Study conducted over 3
generations; BCF for kidney of first
generation.
Study conducted over 3
generations; BCF for liver of
second generation.
Study conducted over 3
generations; BCF for kidney of
second generation.
Exposed for 2 days under static
conditions to 10,000 ug/L; visceral
BCF
Exposed for 2 days under static
conditions to 10,000 ug/L; gill
BCF
-------
Terrestrial AJcity - Lead
CAS No. 7439-92-1
Chemical Name
ead
•
'
lead
lead
lead
lead
lead
Species
chicken
chicken
Japanese
quail
Japanese
quail
Japanese
quail
lambs
Type of
Effect
rep
rep
rep
rep
rep
dev
|
lead I lambs dev
Description
NOAEL
LOAEL
NOAEL
LOAEL
LOAEL
LOAEL
NOAEL
Value
1.56
3.12
0.15
1.53
0.21
4.5
2.3
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
mg/kg-day oral
Exposure
Duration/
Timing
4 weeks
*\ W66KS
5 weeks
5 weeks
5 weeks
6 months
6 months
Reference
Edens and Garllch, 1983
Edens and Garlich, 1983
Edens and Gartich, 1983
Edens and Garlich, 1983
Edens and Garlich, 1983^
Carson etal., 1974
Carson etal., 1974
Comments
•
Egg production significantly
decreased compared to
controls
Egg production significantly
decreased campared to
controls. Birds six weeks old
at start of experiment.
Birds dosed from day of
hatch.
Ewes exposed 5 weeks
before breeding and
throughout gestation.
Learning abilty experiments
run on lambs.
Lambs prenatally exposed to
maternal blood levels
corresponding to 4.5 mg/kg-
day required significantly
more days to leam visual
discrimination problems.
-------
Freshwater Toxicity - Lead
CAS No. 7439-92-1
Chemical Name
lead
ead
lead
lead
lead
/
lead
lead
lead
Species
green frog
adpoles
ainbow trout
rainbow trout
brook trout
smallmouth
bass
fingerlings
aquatic
organisms
fish
daphnids
Type of Effect
behv
dev
dev
dev
dev, beh
chron
chron
chron
Description
NOEC
LOEC
NOEC
LOEC
NOAEL
AWQC
CV
CV
Value
.
1000
7:6
4
119
405
3.2
18.8B
12.26
Units
ug/L
ug/L
"9/L .
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(static/flow
through)
NS
NS
NS
NS
NS
NA
NA
NA
Exposure
Duration/ '
Timing
6 days
19 months
1 9 months
3 generations
90 days
NS
NS
NS
Reference
Taylor el al., 1990
Daviesetal., 1976
Daviesetal., 1976
Holcombe et al , 1976
as cited in Spry and
Wiener. 1991
Coughlan et al.. 1986
as tiled in FWS. 1986
U.S. EPA. 1985
Suteretal., 1992
Suterelal , 1992
Comments
Behavioral effects examined were
preference/avoidance responses and
spontaneous locomotor activities.
Fish developed black tail which can
lead to bending of the spine.
Fish developed black tail which can
lead to bending of the spine.
Fish developed black tail which can
lead to bending of the spine.
No effect on growth or behavior.
Hardness dependent criteria
normalized to 100 mg/L
-------
Terrestrial Toxidty - Lead
CAS No. 7439-92-1
Chemical Name
lead
lead
lead
lead
lead
lead
lead
Species
American
kestrel
American
kestrel
chicken
calves
calves
mouse
mouse
Type of
Effect
systemic
systemic
dev
NS
chron
rep
rep, (et
Description .
NOAEL
LOAEL
PEL
NOEL
PEL
PEL
PEL
Value
1.26
6.3
195.8
3
5
3
20
Units .
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-body
wt.
Exposure
Route (oral,
8.C., I.W.. l.p.,
Infection)
oral
oral
oral
ora[
oral
oral '
i.v. injection
Exposure
Duration/
Timing
5-7 months
5-7 months
4 weeks
3 months
3 years
dose given 3-
5 days after
mating
day 8 of
gestation
Reference
Fransonetal , 1983
Pranson etal., 1983
Franson and Custer, 1982
ZmudsJuet_al.L1983
Zmudski et al., 1983
Clark, 1979
Wide, 1985
Comments
No significant difference (of
liver residue) between the
controls and the 1 0 ppm level
were observed.
Liver residues from birds in
the 50 ppm group were
greater than residues in the
two other treatment groups.
Growth rate suppressed.
No indication that this was a
dose response study,
however no effects were
observed at this dosing
regime.
No indication of a dose '
response study or description
of specific effects;
documented effect was
'chronic toxicity.'
Frequency of pregnancy
reduced.
Smaller litters, increased fetal
deaths.
-------
Terrestrial . .jcity - Lead
CAS No. 7439-92-1
Chemical Name
lead
lead
lead
lead
lead
lead
lead
lead
lead
Species
mouse
rat
rat
rat
rat
rat
rat
rat
American
kestrel
Type ol
Effect
acute
rep
rep
rep
dev
dev
neuro
neuro
dev
Description
LD50
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
NOAEL
LOAEL
NOAEL
Value
890
22
45
0.02
64
0.6
0.002
0.01
6.3
Units
mg/kg-body
wt.
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C., I.V., l.p.,
Injection)
oral
oral (drinking
water)
oral (drinking
water)
oral
gavage in
water
oral
oral
oral
oral
Exposure
Duration/
Timing
NS
60 days
60 days
30 days
gestation
daysl -21;
1/day
dams treated
during
gestation and
lactation
6- 12 months
6- 12 months
6 months
Reference
RTECS, 1994
Chowdhury et at.. 1984 as
cited in ATSDR, 1993
Chowdhury et at., 1984 as
ci'MilATSDR. 1??3
Hilderbrand etal., 1973
Miller etal.. 1982 as cited
In ATSDRL 1993
Reiter etal.. 1975
Krasovskii et al., 1979
Krasovskii et al., 1979
Pattee, 1984
Comments
No significant changes were'
observed at this dose level.
The seminiferous tubular
diameter and spermatic count
were reduced at this dose
level.
Irregular estrus cycles in
females.
A decrease in fetal weight
was observed at this dose
level.
A delay in nervous system
development as seen by a
delay in the righting reflex was
observed at this dose level.
No deviations in functional
state were observed at this
dose level in comparison to
the control group.
Disruption of conditioned
responses and motor activity.
No observable effects on
survival, egg laying, fertility or
eggshell thickness.
i
-------
Terrestrial Toxicity - Lead
CAS No. 7439-92-1
Chemical Name
lead
lead
ead
lead
lead
lead
lead
Species
Rhesus
monkey
Cynmolgus
monkey
Rhesus
monkey
raptor
raptor
American
kestrel
American
kestrel
Type of
Effect
NS
rep
behv
chronic
chronic
dev
dev
•Description
NOEL
PEL
PEL
PEL
PEL
NOAEL
LOAEL
Value
20
5
0.5
9
3
25
125
Units
mg/L
mg/kg-day
mg/kg-diet
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
B.C., I.V., l.p.,
Injection)
oral (drinking
water)
i.m. injection
oral
oral
oral
oral
oral
Exposure
Duration/
Timing
4 weeks
during
pregnancy or
lactation •
4 weeks
30 weeks _
12 weeks
1 0 days
10 days
Reference
Nriagu, 1978 as cited in
FWS, 1988
Tachon etal.,1983
Nriagu, 1978 as cited in
FWS, 1988
Reiser and Temple. 1981
Reiser and Temple, 1981
Hoffman el al.. 1985
Hoffman et al., 1985
Comments
Not a dose response study:
no effects observed.
Abortions and death in
pregnant monkeys.
Abnormal social behavior.
Illness and death associated
with lead toxicosis occured in
44% of experimental birds. In
88% of birds clinical signs of
lead toxicosis were evident.
Vulture used in study showed
clinical signs of lead toxicity
Red-tailed hawks did not.
Growth rate of nestlings
significantly lower than
controls.
-------
Terrestrial xicity - Lead
CAS No; 7439-92-1
Chemical Name
ead
lead
lead
lead
lead
lead
lead
Species
dog
mouse
mouse
sheep
sheep
mouse
mouse
Type of
Effect
chron
dev.rep
dev
t
NS
rep
dev, rep
dev, rep
Description
FEU
NQAEL
PEL
NOEL
PEL
LOAEL
NOAEL
Value
0.32
1,000
5
5
9
25
5
Units
mg/kg-day
mg/L
mg/L
mg/kg-day
mg/kg-day
m9A
mg/L
Exposure
Route (oral,
8.C., I.V., l.p.,
m(ectlon)
oral
oral (drinking
water)
oral (drinking
water)
NS
oral
oral (drinking
water)
oral (drinking
water)
Exposure
Duration/
Timing
NS
9 months
lifetime
1 year
throughout
pregnancy
6-7 weeks
prior to
mating and
continuously
throughout
pregnancy.
6-7 weeks
prior to
mating and
continuously
throughout
pregnancy.
Reference
Demayoetal., 1982
Demayoetal., 1982
Demayoetal., 1982
NRCC, 1973
Forbes and Sanderson.
1978 as cited in FWS,
1988
Kimmelelal , 1980
Kimmel et al . 1980
Comments
Documented effect was a
'chronic toxic level.' ,
Not a dose response study;
No effect on survival or
fertility.
Lowered survival and reduced
longevity.
Not a dose response study;
no adverse effects.
Abortions and death in
pregnant sheep.
Growth retardation, delayed
vaginal opening and some
maternal deaths.
-------
APPENDIX B Lindane - 1
Toxicological Profile for Selected Ecological Receptors
Lindane
CasNo.: 58-89-9
Summary: This profile on lindane summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for
birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), .bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the
freshwater ecosystem were calculated for organic constituents with log Kow between 4 and
6.5. For tile terrestrial ecosystem, these biological uptake measures also include terrestrial
vertebrates and invertebrates (e.g., earthworms). The entire lexicological data base compiled
during this effort is presented at the end of this profile. This profile represents the most
current information and may differ from data presented in the technical support document for
the Hazardous Waste Indentification Rule (HWIR): .Risk Assessment for Human and
Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (€_) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable studies were found for lindane toxicity in mammalian species
associated with the freshwater ecosystem.
Birds: Limited studies were identified concerning the effects of lindane on birds. Ware and
Naber (1961) observed no effect on body weight gain, mortality, clinical symptoms, or egg
production in laying hens that were fed diets containing lindane at 0.01, 0.1, 1.0, or 10
mg/kg-diet for a period of 60 days. Based on these observations, a NOAEL of 10 mg/kg-diet
could be inferred, corresponding to a daily dose of 0.494 mg/kg-day. This daily dose was
based on the body weight of 1.6 kg (U.S. EPA, 19881) and the drived food intake rate of
0.079 mg/kg-day (Nagy, 1987) for mature female chickens. In two experiments that were
part of a study examining egg production, Whitehead el al. (1972a) administered 100 mg/kg
August 1995
-------
APPENDIX B Lindane - 2
Undone to laying hens. The dose corresponded to a daily-dose of 7.0 mg/kg-day. The
conversion was based on the derived food intake rate of 0.112 kg/day (Nagy, 1987) and the
mean body weight of 1.6 kg (U.S. EPA, 19881) for mature female chickens. While egg shell
thickness, egg and yolk weight and hatchability were not significantly affected, after 2 weeks,
egg production decreased by 20% to 30%. In a followup study, Whitehead et al. (1972b)
administered lindane in doses of 10, 100 and 200 mg/kg-diet. The authors found that the
shells of the hens' eggs were not adversely affected by the administration of lindane up to the
200 mg/kg-diet level, but it was noted that egg production was reduced at 100 and 200
mg/kg-diet This resulted in an inferred NOAEL and LOAEL of 10 and 100 mg/kg-diet.
These doses were converted to daily doses of 0.7 mg/kg-day and 7 mg/kg-day using the
derived food intake rate of 0.112 kg/day (Nagy, 1987) and the geometric mean body weight
of 1.6 kg (U.S. EPA, 1988) for mature female chickens. In a study conducted by
Chakravarty et al. (1986, as cited in WHO, 1991), lindane was administered by gavage (99.8
percent in olive oil) to four groups of laying ducks. They were dosed with 20 mg/kg-day
lindane daily, three times a week or twice a week for eight weeks. Groups treated daily and
three times weekly stopped laying eggs, and had drastically reduced clutch sizes when egg
laying resumed. Egg laying capacity was restored to normal following a single injection of
stilbesterol, which suggested that lindane imposed its effects by inducing estradiol
insufficiency.
The NOAEL for reproductive effects from the Whitehead et al. (1972b) study was chosen to
derive the lexicological benchmark because (1) chronic exposures were administered via oral
ingestion, (2) the study focused on longterm reproductive success as a critical endpoint, (3)
the study contained dose response information, and (4) the study contained the lowest toxicity
value for a critical endpoinL The Ware and Naber (1961) and Chakravarty et al. (1986, as
cited in WHO, 1991) studies were not chosen for the derivation of the benchmark primarily
because they did not contain sufficient dose response information. Therefore, the NOAEL of
0.7 mg/kg-day from the Whitehead et al. (1972b) study was chosen for the derivation of a
mammalian benchmark value.
The study value from Whitehead et al. (1972b) study was scaled for species representative of
a freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
(XT*
Benchmark = NOAEL. x L
IKJ
where NOAELj is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). The
principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Since the critical endpoint
selected from the Whitehead et al. (1972b) study was the production of eggs by hens, the
August 1995
-------
APPENDIX B Lindane - 3
mean female body weight of representative species was used in the scaling algorithm to
obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of lindane, as well as
growth or chronic survival. In addition, the data set contained studies which were conducted
over chronic and subchronic durations and during sensitive life stages. For avian species
there were no studies identified with a toxicity value at least an order of magnitude below the
benchmark value. Based on the data set for lindane, the benchmarks developed from the
Whitehead et al. (1972b) study were categorized as adequate.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 8.0E-05 mg/L reported in
the AWQC document for lindane (U.S. EPA, 1986) was selected as the benchmark protective
of fish and aquatic invertebrates. Since the FCV is based on an FCV developed for a
AWQC, it was categorized as adequate.
Aquatic plants: The benchmarks for aquatic plants were either (1) a no observed effects
concentration (NOEQ or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g. duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutwri). For lindane, the
benchmark value presented in Suter and Mabrey (1994) of 5.0E+02 was based on 20%
growth inhibition of Scenedesmus acutus. As described in Section 4.3.6, all benchmarks for
aquatic plants were designated as interim.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^ to determine a protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in
sediment while still protecting the benthic community from harmful effects from chemical
exposure. The FCV, taken from the AWQC, for lindane was used to calculate an EQp
number of 3.38E-01 mg lindane per kg organic carbon. Assuming a mass fraction of organic
carbon for the sediment (f^ of 0.05, the benchmark for the benthic community is 1.69E-02
mg lindane per kg of sediment Because the EQ- number was set using a SCV derived from
AWQC, it was categorized as adequate.
August 1995
-------
APPENDIX B
Lindane • 4
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
80Mie*
mink
hv«r otter
baJdeaole
osprey
groat blue heron
mallard
lesser scaup
spotted sandpiper
herring gul
kingfisher
v«to**»a*^
**t
ID
10
0.54 (a)
0.68 (a)
0.65 (a)
0.77 (a)
0.85 (a)
1.69 (a)
0.79 (a)
1.28 (a)
Study
dptcief
-
-
hen
hen,
hen
hen
hen
hen
hen
hen
Ettect
• •
-
rep
rep
rep
rep
rep
rep
rep
rep
«e*H«r
-
-
0.7
0.7
0.7
0.7
0.7
0.7
0.7
0.7
DMttftetfoA
-
•
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL.
+
, <*
•
•
•
-'
•
-
-
OtfefeatftMwe*
-
Whitehead at al..
1972b
Whitehead et al.,
1972b
Whiteheed et al.,
1972b
Whitehead et al.,
1972b
Whitehead et al.,
1972b
Whitehead et al.,
1972b
Whitehead et al.,
1972b
Whitehead etal..
1972b
'Benchmark Category, a « adequate, p = provisional, i = interim; a "" indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
ID * insufficient data
August 1995
-------
APPENDIX B
Lindane - 5
Table 2. Toxkological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
SfMck*
fish and aquatic
invertebrates
aquatic plants
benlhic community
BwKiunaffc
VUti*
**/t
8.0E-05 (a)
5.0E+02 ug/l (i)
3.4E-01 (a)
Sfwfy
fin^U**
•^^^rar^^Y
aquatic
organisms
aquatic
.plants
aquatic
organisms
to**?**
FCV
CV
FCVxK,,.
Qri0toaf$mtte«
AWQC
Suttr and Mabrey,
1994
AWQC
IL
'Benchmark Category, a = adequate, p = provisional, i = interim; a '"' indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
Toxkological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable studies were found for lindane toxicity in mammalian species
associated with the terrestrial ecosystem.
Birds: As in the freshwater ecosystem, the study by Whitehead et al. (1972b) was used to
calculate the benchmarks for birds in the generic terrestrial ecosystem. The study NOAEL of
10 mg/kg-diet (0.7 mg/kg-day) was scaled for the representative species by using the cross-
species scaling algorithm developed by Opresko et al. (1994). Since Whitehead et al. (1972b)
administered dietary doses of lindane to laying hens the mean female body weights for each
representative species were used in the scaling algorithm to obtain the lexicological
benchmarks. Based on the avian data set for lindane, the benchmarks developed from the
Whitehead et al. (1972b) study were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the LOEC values and then approximating the 10 percentile.
If there were 10 or fewer values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, studies were not identified for benchmark development for lindane.
August 1995
-------
APPENDIX B
Lindane - 6
5o// Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
ttArftftoltftkttftfttikM:
F»^(PPF^^^^^*^WP^W'
SfMcte*
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white- tailed daw
red- tailed hawk
American kesM
Northern
bobwhita
American robin
American
woodcock
plants
toil community
V«ta«*
wgtaHv
10
10
10
ID
ID
ID
ID
0.75 (a)
1 .32 (a)
1.23 (a)
1.47 (a)
1.1 7 (a)
ID •
ID •
SHedy
Spedee
-
-
-
-
-
hen
hen
hen
hen
hen
-
-
ffitljMA
Effect
-
-
-
•
•
rep
"»P
rep
rep
rep
0w*y
V*fcw
waft*
. ..**... '
•
-
-
• -
•
-
0.7
0.7
0.7
0.7
0.7
-
OMttripton
,
-
'
-
- •
-
-
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
•
-
m
-
-
•
-
-
-
•
-
-
firifoJHpfr 1^ffWM^
-
-
-
-
Whitahead et al.,
1972
Whitahead etal.,
1972
Whitahead et al.,
1972
Whitahead et al.,
1972
Whitahead etal..
1972
'Benchmark Category, a «
magnitude or more above the
10 = insufficient data
i, p = provisional, i » interim; a "' indicate* that the benchmark value was an order of
NEL or LEL for other adverse effects.
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
August 1995
-------
APPENDIX B Lindane - 7
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the .value reflects total surface water concentrations. For organic
chemicals with log K^ values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for consituents of concern were .predicted or identified
only constituents with log KDW values above 4.0. However/field bioaccumulation data on
Lindane suggested that this consituent bioaccumulates to a much greater extent than would be
expected using bioaccumulation models. Therefore, the geometric mean of measured BAF/s
identified in Stephan (1993) were used for fish in the limnetic and littoral ecosystems. It
should be noted that the BAF/s were measured for trophic level 4 (TL4) fish and using them
to represent trophic level 3 (TL3) fish may overpredict the actual bioaccumulation in smaller
fish assumed for TL3. The bioconcentration factor for fish was estimated from the Thomann
(1989) model (i.e., log K^ - dissolved BCF/) because: (1) the predicted BCF/ was in close
agreement with the geometric mean of 6 measured BCF/s, (2) the BCF/ was in close
agreement with other predicted BCF/ s based on other methods (i.e., regression equations), (3)
there were no data (e.g., metabolism) to suggest that the log Kow « BCF/1 relationship
deviates for Lindane for water only exposures Gog K,,w = 3.7). As stated in section 5.3.2, the
dissolved bioconcentration factor (BCFjd ) for organic chemicals with log K,,w below 4 was
considered to be equivalent to the total bioconcentration factor (BCF/1) and, therefore,
adjusting the BCF, by the dissolved fraction (fd) was not necessary.
The bioaccumulation factor for earthworms was the geometric mean of measured values cited
in Claborn, et al., (1960) as cited in Kenaga (1980). For terrestrial vertebrates and
invertebrates, the BAFs and BCFs were estimated as described in Section 5.3.5.2.3. Briefly,
the extrapolation method is applied to hydrophobia organic chemicals assuming that the
partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs and
BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data is compared to the BBF for TCDD and that ratio (i.e.,
Lindane BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial vertebrates and
invertebrates, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
August 1995
-------
APPENDIX B Lindane-8
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Lindane - 9
Table 4. Biological Uptake Properties
iwMpfor
limnetic frophic
level 4 fish
limnetic. frophic
level 3 fish
fi«h
littoral frophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
8Cf,**fv«r
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
$*MN»«d«
^fttkl^lwuftf
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
lipid
wtio(*-plant
-------
APPENDIX B Lindane - 10
References
AQUIRE (AQU&tic Toxicity /nformation /JEtrieval Database). 1994. Environmental Research
Laboratory, Office of Research and Development, US. EnvironrrientalProtection Agency,
Duluth, MN.
Chakravarty, S.A. Mandal, and P. Lahiri. 1986. Effect of lindane on clutch size and level of
egg protein in domestic duck. Toxicology. 39: 93-103. As cited in WHO (World Health
Organization), 1991, Lindane, Environmental Health Criteria 124, Geneva, Switzerland..
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Milk. ARS-33-63. U.S. Department of Agriculture. As cited in Kenaga* E.E., 1980,
Correlation of bioconcentration factors of chemicals in aquatic and terrestrial organisms
with their physical and chemical properties, Environmental Sci. Technnol. 14(5):553-556.
Dunachie, J.F. and W. W. Fletcher. 1966. Effect of some insecticides on the hatching rate of
hens' eggs. Nature, 212:1062-1063.
Earl, F.L., E. Miller, and E.J. Van Loon. 1973. Reproductive, teratogenie, and neonatal
effects of some pesticides and related compounds in beagle dogs and minature swine. In.
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Report: A Cross-Species Scaling Factor for Carcinogenic Risk Assessment Based on
Equivalence of mg/kg^/day.
Harrison, D.L., Poole W.S.H., and Mol, J.C.M. 1963. Observations on feeding lindane-
fortified mash to chickens. NZ Vet J. 11 (6): 137-140. As cited in WHO (World Health
Organization), 1991, Undone, Environmental Health Criteria 124, Geneva, Switzerland.
Herbst, M., and G. Bodenstein. 1972. Toxicology of lindane. In Lindane. Edited by E.
Ulman, p. 23. Verlag K. Schillinger, Freiburg, Breisgau.
IARC (International Agency For Research on Cancer). IARC Monographs on the Evaluation
of the Carcinogenic Risk of Chemicals to Humans: Volume 20 Some Halogenated
Hydrocarbons. 1979.
August 1995
-------
APPENDIX B Lindane - 11
Katz, M. 1961. Acute toxicity of some organic insecticides to three species of salmonids
and to the threespine stickleback. Trans. Am. Fish. Soc. 90(3):264-268. As cited in
AQUIRE G4Qt/atic Toxicity /nformation /?£trieval Database), Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Kenaga, E.E. 1980. Correlation of Bioconcentration Factors of chemicals in aquatic and
terrestrial organisms with their physical and chemical properties. Environmental Science
and Technology 14(5):553-556.
Khera, K.S., C. Whalen, G. Trivett, and G. Angers. 1979. Teratogenicity studies on
pesticidal formulations of dimethoate, diuron and lindane in rats. Bull. Environm.
Contain. Toxicol. 22:522-529.
Lutz-Ostertag, Y. 1974. Study over several generations of the effects of lindane on fertility
rate and embryo mortality, hatching, laying, and the weight of eggs and chicks (Fr.).
Arch. Anat. Hist. Embr. Norm. Exp., 57:269-282.
Macek, K.J., C, Hutchinson, and O.B. Cope. 1969. The effects of temperature on the
susceptibility of bluegills and rainbow trout to selected pesticides. Bull. Environ. Contain.
Toxicol. 4(3): 174-183. As cited in AQUIRE (AQUatic Toxicity /nformation /?£trieval
Database), Environmental Research Laboratory, Office of Research and Development,
U.S. Environmental Protection Agency, Duluth, MN.
Macek, K.J., K.S. Buxton, S.K. Derr, J.W. Dean, and S. Sauter. U.S. EPA, 1976. Chronic
toxicity of lindane to selected aquatic invertebrates and fish. EPA-600/3-76-046.
Environmental Resource Laboratory, U.S. Environmental Protection Agency, Washington,
D.C. As cited in AQUIRE (AQt/atic Toxicity /nformation KEtrieval Database),
Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, Duluth, MN.
Macek, K.J., K.S. Buxton, S.K. Derr, J.W. Dean, and S. Sauter. U.S. EPA, 1976. Chronic
toxicity of lindane to selected aquatic invertebrates and fish. EPA-600/3-76-046.
Environmental Resource Laboratory, U.S. Environmental Protection Agency, Washington,
D.C. As cited in Rand, G.M. and S.R. Petrocelli, 1985, Fundamentals of Aquatic
Toxicology, Hemisphere Publishing Corporation, New York.
Macek, K.J., K.S. Buxton, S.K..Derr, J.W. Dean, and S. Sauter. U.S. EPA, 1976. Chronic
toxicity of lindane to selected aquatic invertebrates and fish. EPA-600/3-76-046.
Environmental Resource Laboratory, U.S. Environmental Protection Agency, Washington,
D.C. As cited in WHO (World Health Organization), 1991, Lindane, Environmental
Health Criteria 124, Geneva, Switzerland.
August 1995
-------
APPENDIX B Lindane - 12
Macek, K.J. and W.A. McAllister. 1970. Insecticide susceptibility of some common fish
family representatives. Trans. Am. Fish. Soc. 99(l):20-27
Marliac, J.P. 1964. Toxicity and teratogenic effects of 12 pesticides in the chick embryo.
Federation Proc., 23:105.
McParland, P.J., and R.M. McCracker. 1973. Vet. Rec. 93:369. As cited in Khera, K.S., C.
Whalen, G. Trivett, and G. Angers. 1979. Teratogenicity studies on pesticidal formulations
of dimethoate, diuron and Undone in rats. Bull. Environm. Contain. Toxicol. 22:522-529.
Nagy, K.A. 1897. Field metabolic rate and food requirement scaling in mammals and birds.
Ecological Monographs 57:111-128.
Naishtein, S.Y., and D.L. Leibovich. 1971. Effect of small doses of DDT and lindane and
their mixture on sexual function and embryogenesis in rats, Hyg. Sanit., 36:190-195. As
cited in LARC (International Agency for Research on. Cancer). 1979.1 ARC Monographs
on the Evaluation of the Carcinogenic Risk of Chemicals to Humans.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
Oliver, B.G. and A.L Niimi. 1985. Bioconcentration factors of some halogenated organics
for rainbow trout: Limitations in their use for predictions of environmental residues.
Environ. Sci. Technol. 22:388-397. As cited in U.S. EPA (U.S. Environmental Protection
Agency). 1993, Derivations of Proposed Human Health and Wildlife Biodccumulation
Factors for the Great Lakes Initiative, PB93-154672, Environmental Research
Laboratory, Office of Research and Development, Duluth, MM.
Oliver, B.G. and A.J. Niimi. 1988. Trophodynamic analysis of polychlorinated biphenyl
congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ.
Sci. Technol. 22:388-397. As cited in Parkerton, T.F. J.P. Connolly, R.V. Thomann, and
C.G. Uchrin, 1993, Do aquatic effects or human health end points govern the development
of sediment-quality criteria for nonionic organic chemicals? Environ. Tox. and Chem.
12:507-523.
Opresko, D.M., B.E. Sample, and G.W. Suter II. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1.
Palmer, A.K. A.M. Bottomley, A. N. Worden, H. Frohberg, and A. Bauer. 1978. Effect of
Lindane on Pregnancy in the rabbit and rat. Toxicology, 9:239-247.
Rand, G.M. and S.R. Petrocelli. 1985. Fundamentals of Aquatic Toxicology: Methods and
Applications. Hemisphere Publishing Corporation, New York.
August 1995
-------
APPENDIX B , Lindane - 13
Rogers, J.H., K.L. Kickson, and MJ. DeFoer. 1983. Bioconcentration of lindane and
naphthalene in bluegills (lepomis macrochirus). In: Aquatic Toxicology and Hazard
Assessment: Sixth Symposium, W.E. Bishop, R.D. Cardwell, and B.B. Heidolph, Eds.
ASTM STP 802. American Society for Testing and Materials, Philadelphia, PA. pp. 300-
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of Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative, PB93-154672, Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Smith, S.I., C.W. Webb, and B.L. Reid. 1970. The effect of injection of chlorinated
hydrocarbon pesticides on hatchability of eggs. Toxicology and Applied Pharmacology.
16: 179-185.
Stephan, C.E., 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. Office of Research and Development, U.S.
Environmental Research Laboratory. PB93-154672. Springfield, VA.
Suter n, G.W., M.A. Futrell, and G.A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
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food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connely, and T. F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
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Travis, C.C. and A.D. Arms. 1988. Bioconcentration of organics in beef, milk, and
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U.S. EPA (U.S. Environmental Protection Agency). 1986. Quality Criteria for Water. EPA
440/5-86-001. Environmental Criteria and Assessment Office, Office of Water
Regulations and Standards, Washington, D.C.
August 1995
-------
APPENDIX B Lindane - 14
U.S. EPA (Environmental Protection Agency). 19881. Recommendations for and
Documentation of Biological Values for.Use in Risk Assessment. EPA P338-179874. U.S.
EPA, Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1990. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, D.C. January. As cited in Pierson,
T.K., A.E. Crook, S.M. Beaulieu, P.N. Graham, N.B. Jones, A.M. Reynolds, and G.P.
Vegh, 1994, Development of Human Health Based Exit Criteria for the Hazardous Waste
Identification Project, Phase HI Analysis.
U.S. EPA (U.S. Environmental Protection Agency). 1993. Derivations of Proposed Human
Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative. PB93-
154672. Environmental Research Laboratory, Office of Research and Development,
Duluth, MN.
Ulmann, E. (ed.) 1972. Lindane: Monograph of an Insecticide, Schillinger, Freiburg. As
cited in Palmer, A.K., A.M. Bohomley, A.N. Worden, H.Frohberg, and A. Bauer. 1978.
Effect of lindane on pregnancy in the rabbit and rat. Toxicology. 9:239-247.
Veith, G.D., D.L. DeFoe, and B.V. BergstedL 1979. Measuring and estimating the
bioconcentration factor of chemicals in fish. /. Fish. Fes. Board Can. 36:1040-1048.
Veith, G.D., D.L. DeFoe, and B.V. BergstedL 1979. Measuring and estimating the
bioconcentration factor of chemicals in fish. /. Fish. Fes. Board Can. 36:1040-1048. As
cited in U.S. EPA (U.S. Environmental Protection Agency). 1993, Derivations of
Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative, PB93-154672, Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
Ware, G.W. and E.G. Naber. 1961. Lindane in eggs and chicken tissues. J. Econ. Entomol.
54 (4):675-677.
Whitehead, C.C., A.C. Downing, and R.J. Pettigrew. 1972. The effects of lindane on laying
hens. Br. Poult. Sci; 13:293-299; C.C. Whitehead, J.N. Downie, and J.A. Phillips. 1972b.
BHC not found to reduce the shell quality of hen's eggs. Nature. 239: 411-412.
Whitehead,C.C, J.N. Downie, and J.A. Phillips. 1974. Some characteristics of the egg shells
of quail fed gamma-BHC. Pestic Sci. 5:275-279. As cited in WHO (World Health
Organization), 1991, Lindane, Environmental Health Criteria 124, Geneva, Switzerland.
August 1995
-------
APPENDIX B Lindane - IS
Whitten, B.K. and C.'J. Goodnight. 1966. Toxicity of some common insecticides to
tubificids. /. Water Pollut. Control Fed.'38(2):227-235. As cited in AQUIRE (AQUwc
Toxicity /nformation flEtrieval Database), Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency.-Duluth, MN.
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Geneva, Switzerland.
Will, M. E., and G. W. Suter, II. 1994. Toxicological benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for the U. S. Department of Energy.
August 1995
-------
Terrestrial Toxicity - Lindane
Cas No. 58-89-9
Chemical
Name
lindane
lindane
lindane
lindane
lindane
lindane
lindane
Species
rats
rats
New Zealand
white rabbits
mice
rat
rat
rat
Type of
Effect
rep
ter
ter
rep
hist
hist
ter
Description
NOAEL
NOAEL
NOAEL
PEL
NOAEL
LOAEL
NOAEL
Value
100
15
15
6
25
50
3.32
Units
ppm
mg/kg-day
mg/kg-day
mg/kg-day
ppm
ppm
mg/kg-day
Exposure
Route (oral,
S.C., I.V., l.p.,
Injection)
oral
gavage
gavage
0.5% w/v
aqueous
suspension
oral
oral
gavage
Exposure
Duration/Timing
3-generation
gestation days 6- 16
gestation days 6- 18
gestation days 11-
13 or days 6 to 15
3-generation study
3-generatipn study
gestation days 6- 15
Reference
Palmer et al., 1978a
Palmer etal.,1978b
Palmer eta)., 1978b
• Ulmann, 1972 as
cited in Palmer et
al.. 1978a
Herbst and
Bodenstein, 1972
Herbst and
Bodenstein, 1972
Kheraetal , 1979
Comments
'There were no compound-related
effects on reproduction and no
compound-related teratogenic
effects in any generation.
No compound-related teratogenic
effects in rats were observed.
No compound-related teratogenic
effects in rats were observed.
Mice dosed from days 6 to 15
produced a higher proportion of
undersized young or 'runts'.
Effects were not seen at this dose
level.
Histology of the liver showed a
greater number of enlarged
heptocytes with plasma margination
and vacuolisation.
-------
Terrestrial 1, jity - Lindane
Cas No. 58-89-9
-
Chemical
Name
lindane
lindane
lindane
lindane
lindane
lindane
Species
female rats
rats
pregnant cows
female beagle
dogs
beagle dogs
hens
Type o!
Effect
rep, dvp
path
ter
ter
path
rep
Description
PEL
NOAEL
NOAEL
PEL
NOAEL
NOAEL
Value
0.5
2.5
70
7.5
1.6
0.7
Units
mg/kg-day
mq/kg-day
grams
ma/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
8.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
4 months
2 years
6 to 1 7 weeks, prior
to giving birth
Prom day 5 after
breeding
throughout the
gestation period
2 years
varied
Reference
Naishtein and
Leibovich, 1971 as
cited in (ARC, 1979
Fitzhugh, 1950 as
cited in IRIS, 1992
McPartand and
McCracker, 1973 as
cited in Khera et al.,
1979
Earietal., 1973
Riven etal., 1978
as cited in IRIS,
1992
Whitehead et al ,
1972b
Comments
•produced disturbances of 4he
oestrous cycle, inhibited the animals
capacity for conception and fertility.
lowered the viability of embryos and
delayed their physical development*
Slight liver and kidney damage and
increased liver weights were noted
at this level. .
'All cows had convulsive seizures,
but recovered and produced normal
calves.'
30.5% pups were stillborn
Treatment-related effects were not
noted at this dose.
No significant effects were seen on
egg production at this level.
-------
Terrestrial Toxicity • Lindane
Cas No. 58-89-9
Chemical
Name
lindane
lindane
lindane
lindane
lindane
lindane
•
Species
hens
hens
hens
white Leghorn x
Australorp
chickens
hens
fertile hen eggs
Type of
Effect
rep
rep
rep
path
rep
dvp
Description
LOAEL
PEL
NOAEL
NOEL
NOEL
NOAEL
Value
7
7
10
4
200
2
Units
mg/kg-day
mg/kg-day
mg/kg-diet
mg/kg-diet
mg/kg-diet
mg/egg
Exposure
Route (oral,
s.c., l.v., l.p.,
Injection)
-
oral
oral
oral
oral
oral
injection via
corn carrier
Exposure
Duration/Timing
varied
6 weeks
60 days
27 days
28 weeks
injected either prior
to incubation or
after a 7-day
incubation period
Reference
Whiteheadetal.,
1972b
Whitehead et al.,
1972a
Ware and Naber.
1961
Harrison et al.. 1963
as cited in WHO,
1991
Whitehead et al.,
1974
Smith etal , 1970
Comments
Egg production was reduced 'by
30.1% at this dose level.
A significant decrease in the rate of
egg production was observed at this
longer duration. However, egg shell
thickness, egg and yolk weight, and
hatchability were not significantly
affected.
No effect was observed on body
weight gain, mortality, clinical
symptoms, or egg production at this
highest dose.
No pathological change was
observed in the animals given 4
mg/kq-diet.
Marliac, 1964 injected 5 mg of
. lindane/egg with no pronounced
increase in hatchability.
-------
Terrestrial 1.. -ity - Lindane
Cas No. 58-89-9
Chemical
Name
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
Species
laying ducks
rat
mouse
dog
cat
rabbit
guinea pig
hamster
NS = Not specified
Type of
Effect
rep
acute
acute
acute
acute
acute
acute .
acute
Description
PEL
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
20
76
44
40
25
60
127
360
Units
mg/kg-day
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral.
B.C., I.V., l.p.,
Inlectlon)
gavage
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
daily, 3 times per
week or twice a
week for eight
weeks
NS
NS
NS
NS
NS
NS
NS
Reference
Chakravarty et al ,
1986 as cited WHO,
1991
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
Comments
Groups treated daily and 3 times per
week stopped laying eggs and had
drastically reduced clutch sizes.
-------
Freshwater Toxicity - Lindane
Cas No. 58-89-9
Chemical
Name
f>
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
Species
aquatic
organisms
fathead
minnow
Daphnia
carinata
Oaphnia
magna
Daphnia
magna
Daphnia
magna
Daphnia
magna
Daphnia
pulex
Type of
Effect
chron
chron
immob.
immob.
rep
growth
mort
immob.
Description
AWQC
MATC
EC50
EC50
EC50
NOEC
LC50
EC50
Value
0.08
9.1-23.5
100
516-6442
(1819.7)
340
150
485 - 1790
'(1072)
460
Units
ug/L
ug/L
ug/L
ug/L
ufl/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NA
NA
NA
NA
NA
NA
NA
NA
Exposure
Duration/
Timing
NS
complete life
cycle test
48 hour
48 hour
16 day
16 day
48 hour
48 hour
Reference
IRIS, 1993
Macek et al., !976a as
cited in Rand and
Petrocelli, 1985
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
Comments
Critical life stage end
points: adult; growth
Lindar Page 9
-------
Linda- Page 11
Freshwater Toxicity - Lindane
Cas No. 58-89-9
lindane
lindane
lindane
daphnid
daphnid
daphnid
EC20
rep
rep
11
NOEC
NOEC
11
4.3
11-19
ug/L
ug/L
ug/L
NA
NA
NA
NS
64 days
64 days
Suterelal , 1992
Maceketal., 1976 as
cited in WHO, 1991
Maceketal.. 1976 as
cited in WHO, 1991
NS = Not specified
NA = Not applicable -
-------
Freshwater "i ..city • Lindane
Cas No. 58-89-9
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
Daphnia
pulex
channel.
catfish
bluegill
bluegill
brook trout
brook trout
catfish
fathead
minnow
rainbow trout
striped bass
fish
daphnid
mort
mod
mort
mort
mort
mort
mort
mort
mort
mort
chronic
chronic
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
CV
CV
3800
450
37 - 810
(130.92)
29-31
(29.35)
44.3
26
115
56
30 - 32 (30.9)
7.3 - 400
(53.89)
14.6
14.5
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
NA
NA
NA
NA
NA
NA
NA
NA
. NA
NA
NA
NA
48 hour
96 hour
96 hour
21 day
96 hour
11.0 day
96 hour
96 hour
96 hour
96 hour
NS
NS
AQUIRE. 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE. 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994 -
AQUIRE, 1994
AQUIRE, 1994
Suterelal., 1992
Suteretal , 1992
Lindane - Page 10
-------
Freshwater Toxicity - Lindane
Cas No. 58-89-9
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane •
lindane
Daphnia
pulex
channel
catfish .
bluegill
bluegill
brook trout
brook trout
catfish
fathead
minnow
rainbow trout
striped bass
fish
daphnid
mod
mort
mort
mort
mort
mort
mort
mort
mort
mort
chronic
chronic
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
CV
CV
3800
450
37 - 810
(130.92)
29-31
(29.35)
44.3
26
115
56
30 - 32 (30.9)
7.3 - 400
(53.89)
14.6
14,5
ug/L
ug/L
ug/L
ug/L
uo/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L '
ug/L
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
48 hour
96 hour
96 hour
21 day
96 hour
11.0 day
96 hour
96 hour
96 hour
96 hour
NS
. NS
AQUIRE. 1994
AQUIRE, 1994
AQUIRE, 1994
?
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
Suteretal, 1992
Sulerelal , 1992
-------
Freshwater 1 .city - Lindane
Cas No. 58-89-9
Chemical
Name
lindane •
lindane
' lindane
lindane
lindane
lindane
lindane
lindane
Species
aquatic
organisms
fathead
minnow
Daphnia
carinata
Daphnia
magna
Daphnia
magna
Daphnia
magna
Daphnia
magna
Daphnia
pulex
Type of
Effect
chron
chron
immob.
immob.
rep
growth
mod
immob.
Description
AWQC
MATC
EC50
EC50
EC50
NOEC
LC50
EC50
Value
0.08
9.1-23.5
100
516-6442
(1819.7)
340
150
485- 1790
(1072)
460
Units
ug/L
UQ/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NA
NA
NA
NA
NA
NA
NA
NA
Exposure
Duration/
Timing
NS
complete life
cycle test
48 hour
48 hour
16 day
16 day
48 hour
48 hour
Reference
IRIS, 1993
Maceketal , 1976aas
cited in Rand and
Petrocelli, 1985
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
Comments
Critical life stage end
points: adult; growth
I
-------
Freshwater "t jity - Lindane
Cas No. 58-S9-9
lindane
lindane
lindane
daphnid
daphnid
daphnid
EC20
rep
rep
11.
NOEC
NOEC
11
4.3
11-19
tig/l-
ug/ L
ug/L
NA
NA
NA
NS
64 days
64 days
Suteretal , 1992
Maceketal., 1976 as
cited in WHO, 1991
Maceketal.. 1976 as
cited in WHO, 1991
NS = Not specified
NA = Not applicable
-------
Freshwater Biological Uptake Measures - Lindane
Cas No. 58-89-9
Chemical name
lindane
.lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
lindane
Species
fish
fish
fish
fish
rainbow trout
salmon
bluegill
rainbow trout
brooktrout
fathead
minnow
B-factor
(BCF. BAF.
BMP)
BCF
BCF
BCF
V
BCF
BAF
BAF
BCF
BCF
BCF
BCF
Value
43.87
23.68
158.5
212.8
125
848.5
23 - 45 (30.06)
146 - 374 (234)
51 - 108
(73.45)
284 - 674 (447)
Measured or
predicted (m,p)
P
m
m
m
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Stephan, 1993
Veith et al.. 1979 as cited in
Stephan, 1993
Rogers et al., 1983 as cited in
Stephan, 1993
Oliver and Niimi, 1985 as cited
in Stephan, 1993
Oliver and Niimi, 1985 as cited
in Stephan, 1993
Oliver and Niimi, 1988 as cited
in Stephan, 1993
Macek et al.. 1976 as cited in
AQUIRE, 1994
Vigano et al., 1992 as cited in
AQUIRE, 1994
Macek et al., 1976 as cited in
AQUIRE, 1994
Macek et al., 1976 as cited in
AQUIRE, 1994
Comments
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
Normalized to 1.0% lipid
Normalized to 1 .0% lipid
Normalized to 1 .0% lipid
Adults; 735 day duration
3-day duration; newly-
hatched and early juvenile
lifestages
Yearlings; 261 -day test
15 day old lifestage; 304-
day test
-------
Terrestrial Biological Uptake Measures • Lindane
Cas No. 58-89-9
Chemlcaf .
name
lindane
lindane
lindane
lindane
lindane
Species
cattle
cattle
cattle (beef)
cattle (milk)
plants
B-factor
(BCF, BAF,
BMP)
BCF
-
BCF
BTF
BTF
BCF
Value
0.7
0.4
0.0165
0.0025
0.28
Measured or
predicted
(m.D)
m
m
m
m
P
Units
NS
NS
NS
NS
ug/gDW
plant)/(ug/g soil)
Reference
Claborn, et.al., 1960 as cited
in Kenaga, 1980
Claborn, et.al., 1960 as cited
in Kenaga, 1980
Travis and Arms, 1988
Travis and Arms, 1988
U.S. EPA, 1990
Comments
BTF = Biotransfer
factors
BTF = Biotransfer
factors
NS = Not specified
-------
Freshwater Biological i. .ake Measures - Lindane
Cas No. 58-89-9
Chemical name
lindane
Species
fathead
minnow
B-factor
(BCF, BAF,
BMF)
BCF
Value
180
Measured or
predicted (m.pl
m
Units
NS
Reference
Veilhelal., 1979
Comments
NS = Not specified
-------
APPENDIX B Mercury - 1
lexicological Profile for Selected Ecological Receptors
Mercury
CasNo.: 7439-97-6
/ ^^^
Summary: This profile on mercury summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
information presented in the technical support document for the "Hazardous Waste Identification
Rule (HWIR): Risk Assessment for Human and Ecological Receptors."
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C_) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Two subchronic studies were identified which reported dose-response data for
mammalian wildlife. Rhesus monkeys were exposed to methylmercury chloride by gavage at
doses of 0.05. 0.16 or 0.5 mg/kg-day during gestation days 20 through 30. No signs of
malformative effects were seen at the two lower doses (Dougherty et al. 1974). However, the
highest dose level was maternally toxic and abortient, suggesting a NOAEL of 0.16 mg/kg-day
and a LOAEL of 0.5 mg/kg-day for reproductive effects. Wobeser et al. (1976a and 1976b) fed
adult female mink rations containing methylmercury chloride at doses of 0.16,0.27,0.72, 1,2 and
2.3 mg/kg-day. Groups exposed to doses of 0.27 - 2.3 mg/kg-day exhibited clinical signs of
toxicity. The 0.16 mg/kg-day exposure group did not show clinical evidence of toxicity but did
exhibit pathological alterations of the nervous system. The authors stated that clinical signs of
toxicity in the 0.16 mg/kg-day exposure group would have probably emerged if the experiment
August 1995
-------
. APPENDIX B Mercury-2
had lasted longer. A LOAEL of 0.16 mg/kg-day was inferred for pathological alterations from
this study. - \
The Wobeser et al. (1976a and 1976b) study was not considered suitable for the derivation of a
benchmark value because of the uncertainty surrounding the critical endpoints. Pathological
alterations could impair an individual organism's ability to survive, however, it could not be
inferred that these effects generally impair the sustainability of an entire population. The
Dougherty study (1974) was selected for the derivation of protective benchmarks because it
reports reproductive effects that could impair the sustainability of a wildlife population.
Additionally, this study provides a dose range sufficient to establish a dose-response relationship.
Therefore, the NOAEL of 0.16 mg/kg-day was used to extrapolate a mammalian benchmark
value.
The study value from the Dougherty (1974) study was then scaled for species representative of
a freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994):
( bw
Benchmark = NOAELt x L
V
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the .wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and importable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since Dougherty (1974)
documented reproductive effects from methlylmercury exposure to female rhesus monkeys, the
representative body weights of female species were used in the scaling algorithm to obtain
lexicological benchmarks.
Data were available on reproductive, developmental, growth and survival endpoints for
methylmercury exposure. In addition, the data set contained studies which were conducted over
acute and chronic durations and during sensitive life stages. All identified toxicity values for
mammals were within an order of magnitude of the benchmark value. Therefore, based on the
data set for mercury, the benchmarks developed from Dougherty (1974) were categorized as
adequate.
Birds: Several studies were identified which investigated the effects of methylmercury on avian
species. Ring-necked pheasants were exposed to dietary methylmercury at doses equivalent to
0.18, 0.37, and 0.69 mg/kg-day for 12 weeks (Fimreite, 1970). Reduced hatchability and egg
production as well as increased numbers of shell-less eggs were reported at all dose levels. Based
on these results, a LOAEL of 0.18 mg/kg-day can be inferred for reproductive effects. In another
study by Fimreite (1970, as cited in U.S. EPA, 1993a), leghorn cockeral chicks were exposed to
dietary methylmercury at concentrations of 1.1, 2.1, and 3.2 mg/kg-day for 21 days. A
significant increase in mortality occurred at exposure to 3.2 mg/kg-day while chicks maintained
at 2.1 mg/kg-day exhibited decreases in growth. Although this study reports a NOAEL of 2.1
August 1995
-------
APPENDIX B Mercury-3
mg/kg-day for mortality and a LOAEL of 1.1 for growth, it is unclear as to whether these
exposure levels would affect an entire population's survival. Reproductive effects were seen in
white leghorn laying hens when they were exposed to methylmercury at dietary concentrations
of 4.9 and 9.8 mg/kg-day for an unspecified period of time (Scott, 1977). Both dose levels
severely impacted egg production and weight, fertility of eggs, hatchability of fertile eggs, and
eggshell strength.
In a series of studies carried over three generations, Heinz (1974, 1975, 1976a, 1976b, 1979)
assessed the effects of dietary methylmercury on mallard ducks. Adult mallard ducks given doses
of 0.064 and 0.384 mg/kg-day for up to 2 years were monitored for egg production, hatching
success and hatchling viability. Based on an assessment of percent cracked eggs, egg production
or number of eggs producing normal hatchlings, no significant reproductive effects were observed
in the first generation. However, the survival rate of offspring from the 0.384 mg/kg-day
treatment group was significantly lower. Second generation parents on the 0.064 mg/kg-day diet
exhibited abnormal egg-laying behavior, impaired reproduction and their ducklings had a slowed
growth rate. Third generation hens in the 0.064 mg/kg-day treatment group laid fewer viable
eggs than those in the control group. Behavior tests designed to measure approach response to
maternal calls and avoidance response to a frightening stimulus pooled over three generations
indicate the cumulative effects over three generations were significant at the lowest dose level.
Therefore, a LOAEL of 0.064 mg/kg-day was inferred based on adverse reproductive and
behavioral effects across the three generations of mallard ducks.
Although the studies by Fimreite (1971) and Scott (1970) provide reproductive endpoints in
response to multiple, dietary methylmercury dose levels, the results of the Heinz (1974, 1975,
1976a, 1976b, 1979) multigeneration studies were found to be most appropriate for the estimation
of a benchmark value for avian species. These studies provide reproductive and behavioral
effects due to methylmercury exposure over three generations of mallards. From all the avian
studies identified, Heinz (1974, 1975, 1976a, 1976b, 1979) furnished the most conservative dose
level that could impair the survival and reproductive potential of an avian population. Therefore,
the LOAEL of 0.064 mg/kg-day was used to derive a benchmark value for representative avian
species of the freshwater ecosystem.
The LOAEL value from the Heinz (1974, 1975,1976a, 1976b, 1979) was then scaled for species
representative of a freshwater ecosystem using a cross-species scaling algorithm adapted from
Opresko et al. (1994):
Benchmark = NOAEL, x
\ *v ;•
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since Heinz (1974, 1975,
1976a, 1976b, 1979) documented reproductive effects from methylmercury exposure to both male
August 1995
-------
• • ' • . atf
APPENDIX B Mercury - 4
and female mallards, the body weights of both male and female representative species were used
in the scaling algorithm to obtain lexicological benchmarks.
Data were available on reproductive, developmental, growth and survival endpoints for
methylmercury exposure. In addition, the data set contained studies which were conducted over
acute and chronic durations and during sensitive life stages. Therefore, based on the data set for
mercury, the benchmarks developed from Heinz (1974, 1975, 1976a, 1976b, 1979) were
categorized as adequate.
Fish and aquatic invertebrates: A review of the literature revealed that an AWQC is not
available for mercury. Therefore, the Tier II method described in Section 4.3.5 was used to
calculate a Secondary Chronic Value (SCV) of 1.3E-03 mg/L. Tier II values or SCV were
developed so that aquatic benchmarks could be established for chemicals with data sets that did
not fulfill all the requirements of the National AWQC. Because the benchmark is based on an
SCV, this benchmark was categorized as interim.
Aquatic plants: The lexicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (EC^) for species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). For mercury the
benchmark value was determined to be 5.0 mg/L based on the growth inhibition of Microcystis
aeruginosa. As described in Section 4.3.6, all benchmarks were described as interim.
Benthic community: The mercury benchmark protective of benthic organisms is pending a U.S.
EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Mercury - 5
Table 1. lexicological Benchmarks for Representative Mammals and Bir
Associated with Freshwater Ecosystem
B« 0.015(a)
0.007 (a)
0.012 (a)
&**t
dptdiMk
rhesus
monkey
rhesus
monkey
mallards
mallards
mallards
mallards
mallards
mallards
malards
malards
iftwa
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study WM.
1.60E-01
1.60E-01
6.40E-02
6.40E -02
6.40E-02
6.40E-02
6.40E -02
6.406 -02
6.40E -02
6.40E-02
o-**.
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
8f
-
-
10
10
10
10
10
10
10
10
-, ff '.
Orfcinal Source
Dougherty, 1974
Dougherty, 1974
Heinz. (1974, 1975,
1976a. 1976b and
1979)
Heinz. (1974. 1975.
1976a. 1976b and
1979)
Heinz. (1974, 1975.
1976a. 18760 and
1979)
Heinz. (1974, 1975,
1976a, 1976b and
1979)
Heinz. (1974, 1975,
1976a, 1976b and
1979)
Heinz, (1974, 1975,
1976a. 19766 and
1979)
Heinz, (1974, 1975.
1976a, 1976b and
1979)
Heinz, (1974. 1975,
1976a, 1976b and
1979)
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID - Insufficient Data
August 1995
-------
APPENDIX B
Mercury - 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
8p»d«»
fith and aquatic
invertebrates
aquatic plants
benlhic community
SandMiiivii
V«1(M*
mgfc,
1.3E-03(i*)
5.0 (i)
under review
*****
tip***
aquatic
organisms
aquatic
plants
-
OMOfett*
scv
cv
•
Ortohwl3o»o»
Suter and Mabrey,
1994
Suter and Mabrey,
1994
•
IL
'Benchmark Category, a - adequate, p = provisional, i = interim; a "' indicates that the benchmark value
was an order of magnitude or more above the NEL or LEL for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (G^) for the general terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial
ecosystem.
Mammals: As discussed in the rationale for the freshwater ecosystem, there were two
possible studies from which to estimate a benchmark value. Since no additional studies were
identified, the NOAEL of 0.16 mg/kg-day reported by Dougherty et al. (1974) was used to
calculate benchmark values. The study value was scaled for species in the terrestrial
ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994). Since
Dougherty (1974) documented reproductive effects from exposure to mercury in female
rhesus monkeys, the representative body weights of female species were used in the scaling
algorithm to obtain lexicological benchmarks. Based on the data set for mercury from
Dougherty (1974), the benchmarks developed for the terrestrial ecosystem were categorized as
adequate. .
Birds: Other than the studies discussed for the freshwater ecosystem, no avian toxicity data
were identified. Therefore, the LOAEL of 6.40E-02 reported by Heinz (1974, 1975, 1976a,
1976b, 1979) was chosen to calculate a benchmark value for the representative avian species
in the terrestrial ecosystem. The study value was scaled for species in the terrestrial
ecosystem using a cross- species scaling algorithm adapted from Opresko et al. (1994). Since
Heinz documented reproductive effects from exposure to mercury in both male and female
mallards, the body weights of both male and female representative species were used in the
scaling algorithm to obtain toxicological benchmarks. Based on the data set for mercury from
August 1995
-------
APPENDIX B Mercury - 7
the studies conducted by Heinz, the avian benchmarks developed for the terrestrial ecosystem
were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root lengths. As presented in Will and Suter (1994), phytotoxicity
benchmarks were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation. The
selected benchmark for phytotoxic effects of mercury in soils is 0.3 mg/kg based on
unspecified toxic effects (Will and Suter, 1994). Since the study value selected is the 10th
percentile of more than 10 LOEC values, the terrestrial benchmark for mercury is categorized
as provisional.
Soil Community: For the soil community, the toxicological benchmarks were established based
on methods developed by the Dutch National Institute of Public Health and Environmental
Protection (RIVM). In brief, the RJVM approach estimates a concentration at which the no
observed effect concentration (NOEC) for 95 percent of the species within the community is
not exceeded. A minimum data set was established in which key structural and functional
components of the soil community (e.g., microfauna, mesofauna and macrofauna) were
represented Measurement endpoints included reproductive effects as well as measures of
mortality, growth and survival. The derived mercury benchmark deemed protective of the
soil community is 0.9444 mg/kg. Since the mercury data set contains NOECs and/or LOECs
for at least four of the representative species outlined in the minimum data set, the soil
community benchmark is categorized as provisional.
August 1995
-------
APPENDIX B
Mercury • 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Hepmenmtv*
fKMMi09
deer ITIOUM
short-tailed
throw
meadow vole
Eastern
cottontail
rod fox
raccoon
white-tailed deer
rad- tailed hawk
American Kestrel
Northern
bobowhrta
American robin
American
woodcock
plants
soil community
S«»H*m«rtc
V»tue>
mgf*#4*t
0.75 (a)
0.77 (a)
062 (a)
0.26 (a)
0.20 (a)
0.1 9 (a)
. 0.09 (a)
0.006 (a)
0.011 (a)
0.010 (a)
0.012 (a)
0.010 (a)
0.3 (p)
0.9444 (p)
8)«dy
flpeclee
rhesus
monkey
rhesus
'monkey
rhesus
monkey
rhesus
monkey
rhesus
monkey
rhesus
monkey
rhesus
monkey
malards
mallards
mallards
mallards
mallards
terrestrial
plants
soil
invertebrate
Btaet
rep.
rep
rep
rep
rap
rep
rep
rep
rep
rep
rep
rep
unspeci
-fied
chronic
Study
V«*u»
«0ffcM*
1.6OE-01
1.60E-01
1.60E-01
1.60E-01
1.60E-01
1.60E-01
160E-01
6.40E-02
6.40E-02
6.40E-02
6.40E-02
6.40E-02
0.3
0.9444
9MM$Mfeft
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOEC
NOEC
9f
•
•
•
•
10
10
10
10
10
-
QrfetottScwte*/
Dougherty. 1974
Dougherty, 1974
Dougherty, 1974
Dougherty, 1974
Dougherty. 1974
Dougherty, 1974
Dougherty, 1974
Heinz, (1974, 1975,
1976a, 1976to and
1979)
Heinz, (1974, 1975,
1976a. 1976band
1979)
Heinz, (1974, 1975,
1976a, 19765 and
1979)
Heinz. (1974, 1975,
1976a, 1976b and
1979)
Heinz, (1974. 1975.
1976a. 1976b and
1979)
Kabata-Pendias and
Pendus.1984 as cited
inWiU and Suter, 1994
Aldenberg and Slob.
1993
'Benchmark Category, a = adequate, p = provisional, i = interim; a "
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
indicates that the benchmark value was an order of
August 1995
-------
APPENDIX B Mercury - 9
ID. Biological Uptake Measures
, This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants* Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcnetration facctors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e:, dissolved water concentration equals
total water concentration). The following discussion describes the rationale for selecting the
biological uptake factors and provides the context for interpreting the biological uptake values
presented in Table 4.
Bioaccumulation and bioconcentration factors were identified in the Great Lakes Water
Quality Initiative Technical Support Document for the Procedure to Determine
Bioaccumulation Factors (U.S. EPA, 1995a). This document, generated in support of the
Final Water Quality Criteria for the Great Lakes System; Final Rule (60 FR 15366, March
1995), represents the state-of-the-science at the Agency in terms of mercury bioaccurnulation.
These BAF* and BCF* values developed for the Great Lakes effort were considered
appropriate for development of protective exposure concentrations for aquatic wildlife.
The bioaccurnulation factor for terrestrial vertebrates was the geometric mean of measured
values from several sources (e.g, Borg et al., 1970; Finley et al., 1979; Aulerich et al., 1974).
Insufficient data were identified to establish bioconcentration factors for terrestrial
invertebrates and earthworms. The bioconcentration factor for plants was identified in Baes
et al. (1984) for soil-to-plant uptake.
August 1995
-------
APPENDIX B
Mercury - 10
Table 4. Biological Uptake Properties
•QOtagiMi
receptor
limnetic trophic
level 4 fish
limnetic frophic
level 3 fish
fwh
littoral frophic
level 4 fish
littoral frophic
level 3 fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
•arthworm*
ptanu
BCF.MF.or
8SAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
«pMb«t«4of
mfcol»i'fcQ4y
whoto
whole
lipid
whoto
whote
whoto
whoto-body
1
whoto-plant
«riw
140,000 ( t)
27,900 (t)
11, 358 (t)
1 40,000 (t)
27.900 (t)
22,700
2.1
-
2.0E-03
, 4MMWW -' '
U.S. EPA. 1995* (GLWQI)
U.S. EPA. 1995a (GLWQI)
U.S. EPA. 1995a (GLWQI )
U.S. EPA. 1995a (GLWQI)
U.S. EPA. 1995a (GLWQI)
U.S. EPA, 1995a (GLWQI)
Borg et al., 1970; Fintoy et al..
1979; Aulerich at al., 1974
incufficiant data
insufficient data
Baasatal., 1084
d = raters to disserved surface walar concentration
t meters to total surface walar concentration
August 1995
-------
APPENDIX B Mercury - 11
Reference
Aldenberg, T. and W. Slob. 1993. Confidence limits for hazardous concnetrations based on
logistically distributed NOEC toxicity data. Ecotoxicolgy and Environmental Safety.
25:48-63.
Aulerich, R. J., R. K. Ringer, and S. Iwamoto. 1974. Effects of dietary mercury on mink.
Archives of Environmental Contamination and Toxicology 2:43-51.
Baes, C.F., R.D. Sharp, A.L. Sjoreen, and R.W. Shor. 1984. Review and analysis of
Parameters and Assessing Transport of Environmentally Released Radionuclides Through
Agriculture. Oak Ridge National Laboratory, Oak Ridge, TN.
Borg, K., K. Ernie, E. Hanko, and H. Wanntorp. 1970. Experimental secondary methyl
mercury poisoning in the goshawk {Accipiter g. gentilis L.).
Bull, R. 1976. Methyl mercury and the metabolic responses of brain tissue. EPA-600/1 -76-
013. Water Quality Division, EPA. Cincinnati, OH.
Cember, H. E. H. Curtis, and B. G. Blaylock. 1978. Mercury bioconcentration in fish:
temperature and concentration effects. Environmental Pollution 17:311-319.
Cocking, D., R. Hayes, M. King, MJ Rohrer, R. Thomas, and D. Ward. 1991.
Compaitmentalization of mercury in biotic componenets of terrestrial flood plain
ecosystems adjacent to the South River at Waynesboro, VA. Water, Air, Soil Pollution.
57-58:159-170.
Dougherty, W. J., F. Coulston, and L. Goldberg. 1974. Toxicity of methylmercury in
pregnant rhesus monkeys. Toxicology and Applied Pharmacology 39:138.
Driscoll, C.T., C. Van, C.L. Schofield, R. Munson, J. Holsapple. 1994. The mercury cycle
and fish in the Adirondack Lakes. Environ. Sci. Technol. Vol. 28, No. 3: 136-143.
60 FR 15366. March 23, 1995. Final Water Quality Guidance for the Great Lakes System;
'Final Rule.
Fimreite, N. 1970. Effects of methyl mercury treated feed on the mortality and growth of
leghorn cockerels. Canadian Journal of Animal Science 50:387-389. As cited in
U.S. EPA (U.S. Environmental Protection Agency). 1993. Great Lakes Water Quality
Initiative Criteria Documents for the Protection of Wildlife (PROPOSED) DDT; Mercury;
23,7,8-TCDD; PCBs. EPA-822-R-93-007. Office of Science and Technology, Office of
Water, Washington, DC.
August 1995
-------
.APPENDIX B Mercury - 12
Finley, M. T., W. H. Stickel, and R. E. Christensen. 1979. Mercury residues in tissues of
dead and surviving birds fed methylmercury. Bulletin of Environmental Contamination
and Toxicology 21(1/2): 105-110.
Gentile, J. H., S.M. Gentile and G. Hoffman. 1983. The effects of a chronic mercury
exposure on survival, reproduction and population dymanics of Mysidopsis Bahia.
Environmental Toxicology and Chemistry. Vol 2.:61-68.
Heinz, G. H. 1974. Effects of low dietary of methyl mercury on mallard reproduction.
Bulletin of Environmental Contamination and Toxicology 11:386-392.
Heinz, G. H. 1976. Methylmercury: Second-year feeding effects on mallard reproduction
and duckling behavior. Journal of Wildlife Management 40(1):82-90.
Heinz, G. H. 1976a. Methylmercury: second-generation feeding effects on mallard
reproduction and duckling behavior. Journal of Wildlife Management 40(1):82-90.
Heinz, G. H. 1976b. Methylmercury: second-generation reproductive and behavioral effects
of mallard ducks. Journal of Wildlife Management 40(4):710-715.
Heinz, G. H. 1979. Methylmercury: reproductive and behavioral effects on three
generations of mallard ducks. 1979. Journal of Wildlife Management 43(2):394-401.
Hudson, R. H., R. K. Tucker, and M. A. Haegele. 1984. Handbook of toxicity of pesticides
to wildlife. U.S. Fish and Wildlife Service. Publication 153. 90 pp. As cited in
U.S. Department of the Interior, Fish and Wildlife Service, 1987, Mercury Hazards to
Fish, Wildlife, and Invertebrates: A Synoptic Review, Biological Report 85(1.10).
Khera, K. S. 1979. Teratogenic and genetic effects of mercury toxicity. pp. 501-518. In
J. O. Nriagu (ed.). The Biogeochemistry of Mercury in the Environment. Elsevier/North-
Holland Biomedical Press, New York.
Kucera, E. 1983. Mink and otter as indicators of mercury in Manitoba waters. Canadian
Journal of Zoology 61:2250-2256.
Macleod, J. G, and E. Pessah. 1973. Temperature effects on mercury accumulation, toxicity,
and metabolism rate in rainbow trout (Salmo gairdnerf). Journal of the Fisheries
Research Board of Canada 30:485-492.
/
Mason, R.P., H.M. Spliethoff, A.G Aurilio and H.F. Hammond. 1994. The influence of
redox conditions on the speciation, distribution and mobility of mercury and arsenic in
freshwater lakes. Preprint extended abstract presented before the Division of
Environmental Chemistry American Chemical Society, San Diego, CA, March 13 - 18,
1994. pp. 366 - 369.
August 1995
-------
APPENDIX B N Mercury - 13
McKim, J. M., G. F. "Olson, G. W. Holcombe, and C. P. Hunt 1976. Long-term effects of
methymercuric chloride on three generations of brook trout (Salvelinus fontinalis):
toxicity, accumulation, distribution, and elimination. Journal of Fisheries Research Board
of Canada 33:2726-2739. As cited in Environmental Health Criteria 86. Mercury —
Environmental Aspects, World Health Organization, Geneva, 1989.
Merian, E. 1994. Metals and Aquatic Contamination Workshop. Environ. Sci. Technol.,
Vol. 28, no. 3:144 - 146. "
Nriagu, J. O., editor. 1979. The biogeochemistry of mercury in the environment.
Elsevier/North-Holland Biomedical Press, Amsterdam. Chapter 19.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Organization for Economic Co-operation and Development 1974. Mercury and the
Environment. Studies of Mercury Use, Emission, Biological Impact and Control. OECD,
Paris.
Pinkney, A.E., D.T. Logan, S.R. Jenness, A.L. Birks. 1992. Mercury in Maryland: Sources,
Trends, Power Plant Involvement, and Preliminary Assessment of Ecological Risks.
Power Plant Research Program, Maryland Department of Natural Resources.
Prahalad, A.K. and G. Seenayya. 1988. In situ partitioning and biomagnification of mercury
in industrially polluted Husainsagar Lake, Hyderabad, India. Water, Air, and Soil
Pollution. 39:81 - 8.7.
PTI Environmental Services News. Mercury: The Case for Site-Specific Aquatic Criteria.
March 1994.
Reinert, R. E., L. J. Stone, and W. A. Willford. 1974. Effect of temperature of accumulation
of methylmercuric chloride and p.p'DDT by rainbow trout (Salmo gairdnerf). Journal of
. the Research Fisheries Board of Canada 31:1649-1652. As cited in Environmental
Health Criteria 86. Mercury — Environmental Aspects, World Health Organization,
Geneva, 1989.
Scott, M. L. 1977. Effects of PCBs, DDT and mercury compounds in chickens and Japanese
quail. Federation Proceedings 36:1888-1893. As cited in U.S. Department of Interior,
Fish and Wildlife Service, 1987, Mercury Hazards to Fish, Wildlife, and Invertebrates: A
Synoptic Review, Biological Report 85(1.10).
August 1995
-------
APPENDIX B Mercury - 14
Sheffy, T. B., and J. R. St. Amant. 1982. Mercury burdens in furbearers in Wisconsin.
Journal of Wildlife Management 46:1117-1120.
Stephan, C. E. 1993. Deriviations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research Development, Duluth, MN.
Suter H, G. W., M. A. Futrell, and G. A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects of Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
Suter n, G.W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-
96/RL
Suzuki, T. 1979. Dose-effect and dose-response relationships of mercury and its derivatives.
pp. 399-431. In J. O. Nriagu (ed.). The Biogeochemistry of Mercury in the Environment.
Elsevier/North-Holiand Biomedical Press, New York.
U.S. Department of the Interior. 1970. Mercury in the Environment Geological Survey
Professional Paper 713. U.S. Government Printing Office, Washington, D.C.
U.S. Department of Labor and Occupational Safety and Health Administration. August 1975.
Mercury. From the Job Health Hazards Series. OSHA 2234.
U.S. EPA (Environmental Protection Agency). 1971. Mercurial Pesticided, Man and the
Environment. PB-230 321. Washington, D.C.
U.S. EPA (Environmental Protection Agency). 1976. Environmental Health Criteria I.
World Health Organization. WHO/EHC -01.
U.S. EPA (Environmental Protection Agency). 1980. Ambient Water Quality Criteria for
Mercury. U.S. Environmental Protection Agency Rep. 440/5-80-058. Avail, from NTIS,
5285 Port Royal Road, Springfield, VA 22161. As cited in U.S. Department of the
Interior, Fish and Wildlife Service, 1987, Mercury Hazards to Fish, Wildlife, and
Invertebrates: A Synoptic Review, Biological Report 85(1.10).
U.S. EPA (Environmental Protection Agency). 1985. Ambient Water Quality Criteria for
Mercury. U.S. Environmental Protection Agency, Washington, DC. Publication No.
EPA-440/5-84-026.
August 1995
-------
APPENDIX B Mercury - 15
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1993a. Great Lakes Water Quality
Initiative Criteria Documents for the Protection of Wildlife (PROPOSED) DDT; Mercury;
2,3,7,8-TCDD; PCBs. EPA-822-R-93-007. Office of Science and Technology, Office of
Water, Washington, D.C.
U.S. EPA (Environmental Protection Agency). 1994. Mercury Study Report to Congress Vol.
V: An Ecological Assessment for Anthropogenic Mercury Emissions in the United States
Draft. Office of Air Quality Planning and Standards and Office of Research and
Development.
U.S. EPA (Environmental Protection Agency). 1994a. Great Lakes Water Quality Initiative
Technical Support Document for the Procedure to Determine Bioaccumulation Factors -
July 1994. EPA-822-R-94-002.
U.S. EPA (Environmental Protection Agency). 1995a. Great Lakes Water Quality Initiative
Technical Support Document for the Procedure to Determine Bioaccumulation Factors -
March 1995. EPA-820-B-95-005.
Watras, C.J. and N.S- Bloom. 1992. Mercury and methylmercury in individual zooplankton:
implications for bioaccumulation. Limno. Oceanogr. 37(6): 1313-1318.
Will, M.E. and G.W. Suter D. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
.Wobeser, G. N., N. D. Nielsen, and B. Schiefer. 1976a. Mercury and mink I: The use of
mercury contaminated fish as a food for ranch mink. Canadian Journal of Comparative
Medicine 40:30-33.
Wobeser, G. N., N. D. Nielsen, and B. Schiefer. 1976b. Mercury and mink D: The use of
mercury contaminated fish as a food for ranch mink. Canadian Journal of Comparative
Medicine 40:34-45.
August 1995
-------
Terrestrial Toxicity - Mercury
Cas No. 7439-97-6
Chemical
Name
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
Species
dog
pig
mink
river otter
mink
Rhesus
monkey
Rhesus
monkey
cat
rat
mink
mink
Type of
Effect
rep
rep
mortality
mortality
mortality
rep
rep
rep
behv
mortality
•
mortality
Description
PEL
PEL
PEL
PEL
PEL
NOAEL
LOAEL
PEL
PEL
LOAEL
NOAEL
Value
0:1
0.5
1
>2.0
5
0.16
0.5
250
2,000
0.16
0.27
Units
mg/kg-day
mg/kg-day
mg/kg-diet
mg/kg-diet
mg/kg-diet
mg/kg-day
mg/kg-day
ug/kg-day
ug/kg-diet
mg/kg-day
mg/kq-day
Exposure
Route (oral,
8.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
throughout
pregnancy
throughout
pregnancy
NS
NS
NS
gestation
days 20-30
gestation
days 20-30
gestational
days 10-58
NS
93 days
93 days
Reference
Khera, 1979
Khera,1979
Sherry and St. Amant. 1982
as cited in FWS, 1987
Kucera, 1983 as cited in
FWS, 1987
Sherry and St. Amant, 1982
as cited in PWS, 1987
Dougherty et al., 1974
Dougherty et al., 1974
Khera ,1979
Suzuki, 1979 as cited in
FWS, 1987
Wobeser et. al., I976a
Wobeser et. al.. 1976a
Comments
High incidence of stillbirths
High incidence of stillbirths
Fatal to 100% in 2 months
Fatal
All dead in 30-37 days
No measurable effects on
reproduction
Maternally toxic and abortient
Increased incidence of
anomalous fetuses
Adverse behavioral changes in
offspring
-------
Terrestrial Toxicity - Mercury
Cas No. 7439-97-6
Chemical
Name
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
mercury
Species
dog
pig
mink
river otter
mink
Rhesus
monkey
Rhesus
monkey
cat
rat
mink
mink
Type of
Effect
rep
rep
mortality
mortality
mortality
rep
rep
rep
behv
mortality
mortality
Description
PEL
PEL
PEL
PEL
PEL _
NOAEL
LOAEL
PEL
PEL
LOAEL
NOAEL
Value
0.1
0.5
1
>2.0
5
0.16
0.5
250
2,000
0.16
0.27
Units
mg/kg-day
mg/kg-day
mg/kg-diet
mg/kg-diet
mg/kg-diet
mg/kg-day
mg/kg-day
ug/kg-day
ug/Vg-diet
mg/kg-day
mq/kq-day
Exposure
Route (oral,
S.C.. I.V.. l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
throughout
pregnancy
throughout
pregnancy
NS
NS
NS
gestation
days 20-30
gestation
days 20-30
gestational
days 10-58
NS
93 days
93 days
Reference
Khera, 1979
Khera,1979.
Shefty and St. Amant, 1982
as cited in FWS, 1987
Kucera, 1983 as cited in
FWS, 1987
Sheffy and St. Amant. 1982
as cited in FWS, 1987
Dougherty et al, 1974
Dougherty et al., 1974
Khera , 1979
S.uzukl, 1979 as cited in
FWS, 1987
Wobeseret. al., 1976a
Wobeser et. al., 1976a
Comments
High incidence of stillbirths
High incidence of stillbirths
Fatal to 100% in 2 months
Fatal
All dead in 30-37 days
No measurable effects on
reproduction
Maternally toxic and abortient
Increased incidence of
anomalous fetuses
Adverse behavioral changes in
offspring
-------
Terrestrial 1 ity - Mercury
Cos No. 7439-97-6
Chemical
Name
mercury
mercury
mercury
mercury
mercury
mercury
Species
mink
ring-necked
pheasants
leghorn
cockeral
chicks
leghorn
cockeral
chicks
white leghorn
hens
mallard
ducks
Type of
Effect
mortality
rep
mortality
dvp
rep
rep, behv
Description
NOAEL
LOAEL
NOAEL
LOAEL
LOAEL
LOAEL
Value
0.05
0.18
2.1
1.1
4.9
0.064
Units
mg/kg -day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C., I.V., l.p..
Injection)
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
145 days
1 2 weeks
21 days
21 days
NS
3 generations
Reference
Wobeseretal., 1976b
Fimreite, 1971 as cited in
U.S. EPA, 1993a
Fimreite, 1970 as cited in
U.S. EPA, 1993a
Fimreite, 1970 as cited in
U.S. EPA, 1993a
Scon. 1977 as cited in U.S.
EPA, 1993a
Heinz, 1974, 1976a,
19766.1979
Comments
.
Study conducted over 3
generations
NS = Not Specified
-------
Terrestrial 1 ,ity - Mercury
Cas No. 7439-97:6
Chemlce.1
Name
mercury
mercury
mercury
mercury
mercury
mercury
Spades
mink
ring-necked
pheasants
leghorn
cockeral
chicks
leghorn
cockeral
chicks
white leghom
hens •
mallard
ducks
Type of
Effect
mortality
rep
mortality
dvp
rep
reo. behv
Description
NOAEL
tOAEL
NOAEL
LOAEL
LOAEL
LOAEL
Value
0.05
0.18
2.1
1.1
4.9
0.064
Units
mg/kg -day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/ka-day
Exposure
Route (oral,
B.C., I.V., I. p.,
Inlectlon)
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing .
145 days
12 weeks
21 days
21 days
NS
3 generations
Reference
Wobeser et al., 1976b
Fimreite, 1971 as cited in
U.S. EPA. 1993a
Fimreite, 1970 as cited in
U.S. EPA, 1993a
Fimreite, 1 970 as cited in
U.S. EPA, 1993a
Scott, 1977 as cited in U.S.
EPA, 1993a
Heinz, 1974, 1976a,
1976b,1979
Comments
•
•
Study conducted over 3
qenerations
NS = Not Specified
-------
Freshwater Toxicity - Mercury
Cas No. 7439-97-6
Chemical
Name
mercury
(organic)
mercury
(organic^
mercury
(inorganic)
mercury
(inorganic)
mercury
(inorganic)
mercury
(inorganic)
mercury
(inorganic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
Species
daphnid
brook trout
aquatic
organisms
fish
daphnid
fish
daphnid
brook trout
aquatic
organisms
fish
daphnid
fish
daphnid
Type of
Effect
acute
acute
chronic
chronic
chronic
chronic
chronic
rep
chronic
chronic
chronic
chronic
chronic
Description
LC50
LC50
NAWQ
CV
CV
EC20
EC20
CV
NAWQC
CV
CV
EC20
EC20
Value
0.9-3.2
65
t
0.012
<0.23
0.96
0.87
0.87
0.29-0.93
0.0003
0.52
<0.04
<0.03
0.87
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ufl/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
lifetime
96 hours
NS
NS
NS
NS
NS
144 weeks
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1980 as cited
inFWS, 1987
U.S. EPA, 1980 as cited
inFWS, 1987
U.S. EPA, 1985
Suteretal., 1992
Suteretal., 1992
Suteretal., 1992
Suteretal., 1992
McKimetal., 1976 as
cited in WHO, 1989
U.S. EPA, 1985
Suteretal., 1992
Suteretal., 1992
Suteretal., 1992
Suteretal., 1992
Comments
Hardness=45 mg
CaCO3/L, ph= 7.5
Estimate
NS = Not Specified
-------
Terrestrial Biological Uptake Measures - Mercury
Cos No. 7429-97-6 .
Chemical
Name
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury
(mercuric
chloride)
mercury
(mercuric
chloride)
mercury
(organic)
mercury
(organic)
Species
chicken
chicken
mallard duck
mallard duck
red-winged
blackbird
red-winged
blackbird
mink
mink
mink
mink
B-factor.
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
1.25
5
2.1
2.2
2.3
2.1
0.3
3.2
11.1
7.4
Measured
or
Predicted
(m,P)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Borgelal., 1970
Borgetal., 1970
Borgetal., 1970
Borgetal., 1970
Finley etal., 1979
Finleyetal., 1979
Aulerich etal., 1974
Aulerich etal, 1974
Aulerich et al., 1974
Aulerich etal., 1974
Comments
Exposure through diet for 35-42
days to 8 mg/kg; kidney BCF.
Exposure through diet for 35-42
days to 8 mg/kg; liver BCF.
Exposure through diet for 14 days
to 8 mg/kg; liver BCF.
Exposure through diet for 14 days
to 8 mg/kg; kidney BCF.
Exposure through diet for 1 1 days
to 40 mg/kg; liver BCF.
Exposure through diet for 1 1 days
to 40 mg/kg; kidney BCF.
Exposure through diet for 10 days
to 135 mg/Kg; liver BCF.
Exposure through diet for 10 days
to 135 mg/kfl; kidne^BCF.
Exposure through diet for 5 days to
32 mg/kg; liver BCF.
Exposure through diet for 5 days to
32 mg/kg; kidney BCF.
-------
Freshwater Biological jke Measures - Mercury
Cas No. 7439-97-6
Chemical
Name
mecury
(organic)
mecury
(organic)
mecury
(organic)
mecury
(organic)
mecury
(organic)
mecury
(organic)
mercury
mercury
mercury
mercury
(mercuric
chloride)
mercury
(mercuric
chloride)
mercury
(mercuric
chloride)
Species
rainbow trout
rainbow trout
rainbow trout
bluegill
bluegill
bluegill
fish
fish
fish
rainbow trout
rainbow trout
rainbow trout
B-factor
(BCF. BAF.
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BAF
BAF
BCF
BCF
BCF
Value
4,525.00
6,628.00
8,033.00
222.00
1,138.00
2,454.00
13,044.00
130,440.00
60,524.00
5.00
12.00
26.00
Measured
or
Predicted
(m,p)
m
m
m
P
P
P
P
P
P
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Reinert et al., 1974 as cited
in WHO, 1989
Reinert et al., 1974 as cited
in WHO, 1989
Reinert et al., 1974 as cited
in WHO, 1989
Cemberetal., 1978
Cemberetal., 1978
Cemberetal., 1978
Stephan, 1993
Stephan, 1993
Stephan, 1993
MacLeod & Pessah, 1973
MacLeod & Pessah, 1973
MacLeod & Pessah. 1973
Comments
Exposed to .263 ug/L for 84 days;
whole body BCF.
Exposed to .258 ug/L for 84 days;
whole body BCF.
Exposed to .244 ug/L for 84 days;
whole body BCF.
Exposed to .5 ug/L for 28.6 days;
whole body BCF.
Exposed to .5 ug/L for 28.6 days;
whole body BCF.
Exposed to .5 ug/L for 28.6 days;
whole body BCF.
Assumes 85.3% of total mercury in fish
is methylmercury.
Assumes in this case, the HHBAF and
WLBAF are equal.
WLBAF assuming an FCM of 4.64 for
trophic level 3.
Exposed to 50 ug/L for 4 days; BCF
from muscle, skin and bone.
Exposed to 50 ug/L for 4 days; BCF
from muscle, skin and bone.
Exposed to 50 ug/L for 4 days; BCF
from muscle, skin and bone.
NS = Not Specified
-------
Terrestrial Biological L. .Ke Measures - Mercury
Cas No. 7439-97-6
Chemical
Name
mercury
(organic)
mercury
(organic)
mercury
(organic)
mercury .
(organic)
mercury
Species
cowbird
cowbird
qrackle
grackle
plant
B-lactor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
Value
1.7
1.5
1.3
1.1
0.005
Measured
or
Predicted
(m,p)
NS
NS
NS
NS
P
units
NS
NS
NS
NS
(ug/g DW
plant)/(ug/g
soil)
Reference
Finley etal. j 1979
Finleyetal., 1979
Finley etal., 1979
Finleyetal., 1979
U.S. EPA, 1990e
Comments
Exposure through diet for 1 1 days
to 40 mg/kg; liver BCF.
Exposure through diet for 1 1 days
to 40 mg/kg; kidney BCF.
Exposure through diet for 1 1 days
to 40 mg/kg; liver BCF.
Exposure through diet for 1 1 days
to 40 mg/kg; kidney BCF.
NS = Not Specified
-------
APPENDIX B Methoxychlor - 1
Toxicological Profile for Selected Ecological Receptors
Methoxychlor
Cas No.: 72-43-5
Summary: This profile on methoxychlor summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire toxicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from data presented in the technical
support document for the Hazardous Waste Indentification Rule (HWIR): Risk Assessment for
Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
.contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found which reported dose-
response data for mammalian wildlife. However, toxicological studies involving v
methoxychlor exposure to mammals have been conducted using laboratory rats and mice. In
a chronic study, Khera et al. (1978) administered 50, 100, 200, and 400 mg/kg-day of
methoxychlor (suspended in corn oil) to female Wistar rats by gavage on days 6-15 of
gestation. The authors observed fetotoxicity at both 200 and 400 mg/kg-day. Effects
included significant decreases in fetal weight, number of line fetuses per litter and increased
incidences of resorption and malformations. There was also a dose-related increase in the
incidence of wavy ribs at 100, 200, and 400 mg/kg-day. Based on these observations, a
NOAEL of 100 mg/kg-day was inferred for the fetotoxic effects and a NOAEL of 50 mg/kg-
day for the teratogenic effect of wavy ribs. In a reproductive study, Bal (1984) exposed rats
August 1995
-------
APPENDIX B Methoxychlor - 2
to 100 and 200 mg/kg-day methoxychlor via oral gavage. Male rats were exposed for 70
days and female rats were exposed for 14 days. The author observed that both
spermatogenesis in the males and folliculogenesis in the females was inhibited at both dose
levels. This lead to an inferred LOAEL of 100 mg/kg-day for these reproductive effects.
Gellert and Wilson (1979) administered 30 mg/kg-day methyoxyclor by oral gavage to
pregnant female rats for seven days to examine the effects of the chemical on the
reproductive system of male and female offspring. There were no effects observed which led
to an inferred NOAEL of 30 mg/kg-day.
The NOAEL for fetotoxic effects from the Khera et al. (1978) study was chosen to derive the
lexicological benchmark because (1) chronic exposures were administered via oral intubation,
(2) the study focused on effects that could have negative implications on longterm
reproductive success and (3) the study contained sufficient dose response information. The
NOAEL inferred for the teratogenic effect was not used because it would be difficult to
predict if the relatively minor incidence of wavy ribs observed in the study would have any
effect on the longterm viability of the population. The Bal (1984) and Gellert and Wilson
(1979) studies were not chosen for the derivation of the benchmark primarily because they
did not contain sufficient dose response information. Therefore, the NOAEL of 100 mg/kg-
day from the Khera et al. (1978) study was chosen for the derivation of a mammalian
benchmark value.
The study value from Khera et al. (1978) was scaled for species representative of a freshwater
ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994):
( bw
Benchmark = NOAEL. x
l
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and importable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
critical endpoint selected from the Khera et al. (1978) study was the toxicity of the fetus in
female rats and the female rats were dosed individually during gestation, the mean female
body weight of representative species was used in the scaling algorithm to obtain the
lexicological benchmarks.
Data were available on the reproductive and developmental effects of methoxychlor, as well
as growth or chronic survival. In addition, the data set contained studies which were
conducted over chronic and subchronic durations and during sensitive life stages. Based on
the data set for methoxychlor, the benchmarks developed from the Khera et al. (1978) study
were categorized as adequate.
Birds: No suitable studies were found for methoxychlor toxicity in avian species associated
with the freshwater ecosystem.
August 1995
-------
APPENDIX B Methoxychlor - 3
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 3.0E-05 mg/L reported in
the AWQC document for methoxychlor (U.S. EPA, 1980) was selected as the benchmark
protective of fish and aquatic invertebrates. Since the FCV is based on an FCV developed for
an AWQC, it was categorized as adequate.
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g. duckweed) or (2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). For
methoxychlor there was insufficient data for the development of a benchmark value.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^ to determine a protective sediment concentration
(Stephan, 1993). The EQp number is the chemical concentration that may be present in
sediment while still protecting the benthic community from harmful effects from chemical
exposure. The FCV, taken from the ambient water quality criteria, for methoxychlor was
used to calculate an EQp number of 8.52E-01 mg methoxychlor per kg organic carbon.
Assuming a mass fraction of organic carbon for the sediment (f^ of 0.05, the benchmark for
the benthic community is 4.26E-02 mg methoxychlor per kg of sediment. Because the EQp
number was set using a FCV derived from ambient water quality criteria, it was categorized
as adequate.
August 1995
-------
APPENDIX B
Methoxychlor - 4
Table 1. ToxicologicaJ Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
BnOT»eowfa»
fijMCiM
mink
river otter
baldeagto
osprey
graat blue heron
malard
lesser scaup
spotted sandpiper
herring guN
kingfisher
8**njfeflxrii
V4*M»«i9ft9>
day
71.83
39.99
10
ID
ID
ID
ID
ID
ID
ID
9*wtf
ftMclfffr
rat
rat
-
-
'
-
-
-
£Jf*d
feto
fato
-
-
-
-
•
.-
•
ittiidTy Iffcfti^
«**»***
100
100
-
-
.
-
-
DMMfnjpWM
NOAEL
NOAEL
-
-
• •
-
- ' -
' • •
-
*F
•»
-
-
-
-
-
-
-
•
Origin** *OUK»
Kh«ra0tal.. 1978
Krwraatal., 1978
-
•
-
-
-
-
*B0nchnwf<( Category, a > adequate, p * provisional, i - interim; a "' indicate* that the benchmark value was
an order of magnitude or more above the NEL or LEL tor other adverse effects.
ID - insufficient data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ftepr*e«tialfo*
Sp*6**»
fish andaquabc
invertebrates
aquatic plants
bentuc community
Benchmark
¥•>»•*
mgtL
3.0E-OS (a)
ID
4.3E-02 (a)'
StMdy
tto^ids*
aquatic
organisms
aquatic
organisms
P*«ortpao»
FCV
-
FCV
QriQBMtdoWs* '
AWQC
-
AWQC
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates (hat the benchmark
value was an order of magnitude or more above (he NEL or LEL for other adverse effects.
ID = insufficient data
August 1995
-------
APPENDIX B Methoxychlor - 5
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion , no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to methoxychlor.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Khera et al., 1978) was used to derive the methoxychlor benchmark for mammalian
species representative of terrestrial ecosystems. The study NOAEL of 100 mg/kg-day was
scaled for species in the terrestrial ecosystem using a cross-species scaling algorithm
developed by Opresko et al. (1994). Since the critical endpoint selected from the Khera et al.
(1978) study was the toxicity of the fetus in female rats and the female rats were dosed
individually during gestation, the mean female body weight of representative species was used
in the scaling algorithm to obtain the toxicological benchmarks.
Based on the data set for endrin the benchmarks developed from the Kavlock et al. (1981)
study were categorized as adequate.
Birds: No suitable studies were found for methoxychlor toxicity in avian species associated
with the terrestrial ecosystem.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the LOEC values and then approximating the 10 percentile.
If there were 10 or fewer values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, studies were not identified for benchmark development for methoxychlor.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Methoxychlor • 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
WW^W^WWWwPwW"
Specie*
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
whita-tailad daw
red- tailed hawk
American kestrel
Northern
bobwhile
American robin
American
woodcock
plants
soil community
Value* !
I*****
177.15
182.14
148.00
62.53
46.40
44.66
22.27
ID
ID
ID
ID
ID
ID
ID
Study "
ftp er tee
rat
rat
rat
rat
rat
rat
rat
•
•
-
Bfret
feto
feto
feto
feto
feto
feto
feto
-
-
-
-
-
-
«M*
Vrtue
V
100
100
100
100
100
100
100
• . v
•
-
-
-
-
+
Peectfrtea %
s ' x^
NOAEL
NOAEL
NOAEL
. NOAEL
NOAEL
NOAEL
NOAEL
• .
•
.
-
-
*
-
-
-
-
•
-
-
' •
-
-
-
-.
-
.'T-
Kheraetal., 1978
Kheraetal., 1978
Kheraetal.. 1978
Kheraetal., 1978
Kheraetal., 1978
Kheraetal., 1978
Kheraetal., 1978
.
-
-'
'Benchmark Category, a » adequate, p = provisional, i = interim, a "" indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = insufficient data
August 1995
-------
APPENDIX B Methoxychlor - 7
III. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only),, aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants! Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for consituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for methoxychlor.
The bioconcentration factor for fish was also estimated from the Thomann models (i.e., log
Kow ~ dissolved BCF/) and multiplied by the dissolved fraction (f
-------
APPENDIX B Methoxychlor - 8
Further, the method assumes that the BAFs and BCFs for terrestrial wildlife developed for
2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial Wildlife from TCDD and
TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient quality to serve as the standard.
The beef biotransfer factor (BBFs) for a chemical lacking measured data is compared to the
BBF for TCDD and that ratio (i.e., methoxychlor BBF/TCDD BBF) is multiplied by the
TCDD standard for terrestrial vertebrates, invertebrates, and earthworms, respectively. For
hydrophobic organic constituents, the bioconcentration factor for plants was estimated as
described in Section 6.6.1 for above ground leafy vegetables and forage grasses. The BCF is
based on route-to-leaf translocation, direct deposition on leaves and grasses, and uptake into
the plant through air diffusion.
August 1995
-------
APPENDIX B
Methoxychlor • 9
Table 4. Biological Uptake Properties
«cetoetart
t*»p4«r
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral frophic
leveUKsh
littoral trophic
level 3 fish
littoral frophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
pUnU
8Cf.BAF,or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
lfpkHn»>tt or
wtw4»*
-------
APPENDIX B Methoxychlor - 10
References
AQUIRE (AQU&tic Toxicity /nformation flEtrieval Database). 1994. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Bal, H. S. 1984. Efffect of methoxychlor on reproductive system of the rat. Proceedings of
the Society for Experimental Biology and Medicine. 176:187-196.
Bal, H. S. and P. Mungkornkam. 1978. Chronic toxicity effects of methoxychlor on the
reproductive system of the rat. Proceedings of the First International Congress on
Toxicology.
57 FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogenic Risk Assessment Based on
Equivalence of mg/kg3/4/day!
Gardner, D.R., J.R. Bailey et al. 1975. Methoxychlor, It's effect on Evironmental Quality,
ISSN. 0316-0114 Natl Res Council Canada No. 14102.
Gellert, R. J., and C. Wilson. 1979. Reproductive function of rats exposed prenataTly to
pesticides and polychlorinated biphenyls (PCB). Environmental Research. 18:437-443.
Goldman, J. M., R. L. Cooper, G. L. Rehnberg, J. F. Hein, W. K. McElroy, and L. E. Gray,
Jr. 1986. Effects of low subchronic doses of methoxychlor on the rat hypothalmi-pituitary
reproductive axis. Toxicology and Applied Pharmacology. 86:474-483.
Gray, L. E., Jr., J. Ostby, R. Sigmon, J. Ferrell, G. Rehnberg, R. Under, R. Cooper, J.
Goldman, and J. Laskey. 1988. The development of a protocol to assess reproductive
effects of toxicants in the rat. Reproductive Toxicology. 2:281-287.
Gray, L. E., Jr., J. Ostby, R. Sigmon, J. Ferrell, G. Rehnberg, R. Under, R. Cooper, J.
Goldman, V. Slott, and J. Laskey. A dose-response analysis of methoxychlor-induced
alterations of reproductive development and function in the rat. Fundamental and Applied
Toxicology. 12:92-108.
Hansch, C. and A.J. Leo. 1985. Medchem Project Issue No. 26. Claremont, CA: Pomona
College.
Heming, T..A-, A. Sharma, and Y. Kumar. 1989. Time-toxicity relationships in fish exposed to
the organochorine pesticide methoxychlor. Environ. Toxicol. Chem. 8(10):923-932.
August 1995
-------
APPENDIX B Methoxychlor - 11
Howard, P.H. 1991. Handbook of Environmental Fate and Exposure Data for Organic
Chemicals. Volume III: Pesticides. Lewis Publishers. Chelsea, Michigan.
IARC (International Agency for Research of Cancer). 1979. IARC Monographs on the
Evaluation of the Carcinogenic Risk of Chemicals to Humans, Volume 20, Methoxychlor.
Khera, K.S., C. Whalen, and G. Trivett. 1978. Teratogenicity studies on linuron, malathion,
and methoxychlor in rats. Toxicology and Applied Pharmacology. 45:435-444.
Macek, K.J., C. Hutchinson, and O.P. Cope. 1969. The effects of temperature on the
susceptibility of bluegills and rainbow trout to selected pesticides. Bull. Environ. Contain.
Toxicol. 4(3):174-183.
Macklin, A.W., and W.E. Ribelin. 1971. The relation of pesticides to abortion in dairy cattle.
/. Am. vet. med. Assoc. 159:1743-1748.
Martinez, E. M., and W.J. Swartz. 1992. Effects of methoxychlor on the reproductive system
of the adult female mouse: 2. ultrastructural observations. Reproductive Toxicology. 6:93-
98.
Opresko, D. M., B. E. Sample, and G. W. Suter. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1.
Paris, D.F., and D.L. Lewis, 1973. Res Rev 45:95.
Reuber, M. D. 1979. Carcinomas of the liver in Osbome-Mendel rats ingesting
methoxychlor. Life Sciences. 24:1367-1372.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
. Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN, PB93-154672.
Suter n, G.W., and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-
96/R1.
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R. V., J. P. Connely, and T. F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry. 11:615-629.
August 1995
-------
APPENDIX B Methoxychlor - 12
Travis, C. C. and A. D. Arms. 1988. Bioconcentration of organics in beef, milk, and
vegetation. Environmental Science and Technology. 22:271-274.
Trutter, J. 1986. Rabbit teratology study with methoxychlor, technical grade: Final Report:
Project No. 2298-100. Hazleton Laboratories America, Inc. pp!35.
U.S. EPA (Environmental Protection Agency). 1980. Water Quality Criteria Document for
Methoxychlor. 46FR40919.
U.S. EPA (Environmental Protection Agency). 19881. Recommendations for and »
Documentation of Biological Values for Use in Risk Assessment. EPA P338-179874. U.S.
EPA, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1993h. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
Office of Science and Technology, Office of Water, Washington, D.C.
U.S. EPA (Environmental Protection Agency). 1993L Interim Report on Data and Methods
for Assessment of 2,3,7,8-Tetrachlorodibenzo-o-dioxin Risks to Aquatic Life and
Associated Wildlife. EPA/600/R-93/055. Office of Research and Development,
Washington, DC.
U.S. EPA. (Environmental Protection Agency). 1994. Integrated Risk Information System.
July.
Veith, G.D. et al. 1979. / Fish Res Board Can 36:1040-8.
Will, M.E., and G.W. Suter, II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Freshwater Biological Uptake Measures • Methoxychlor
Cas No. 72-43-5
Chemical name
methoxychlor
methoxychlor
methoxychlor
methoxychlor
methoxychlor
Species
fish
fish
fish
common
mirror colored
carp
fathead
minnow
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
Value
467
42
1092
70700
8300
Measured or
predicted
(m.p)
P
m
m
m
m
Units (LAg.
NS, other)
NS
NS
NS
NS
NS
Reference
U.S. EPA. 1993
Parrish et al., 1977 as cited
in U.S. EPA, 1993
Veith et al., 1979 as cited in
U.S. EPA, 1993
Lakota et al., 1978 as cited
in AQUIRE, 1994
Veith et al., 1979 as cited in
AQUIRE, 1994
Comments
1 .0% lipid
1 .0% lipid
1.0% lipid
6 months old, 30-40 grams;
30-day test
adult; 32-day test
NS = Not specified
-------
Freshwater Tox y - Methoxychlor
Cas No. 72-43-5
Chemical
Name
Methoxychlor
Methoxychlor
Methoxychlor
Methoxychlor
Methoxychlor
Species
bluegill
brook trout
rainbow trout
striped bass
fathead
minnow
Type of
Effect
mort
mort
mor
mort
mor
Description
LC50
LC50
LC50
LC50
LC50
Value
40.0 - 75
(50.3)
40
9.36 - 62.6
(33.50)
3.3
7.5 - 35
(1 1 .36)
Units
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NA
NA
NA
NA
NA
Exposure
Duration/
Timing
96 hour
1day
96 hour
96 hour
96 hour
Reference
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE. 1994
AQUIRE, 1994
Comments
NA = Not applicable
-------
Terrestrial Toxicity - Methoxychlor
Cas No. 72-43-5
Chemical
Name
melhoxychlor
methoxychlor
methoxychlor
methoxychlor
methoxychlor
methoxychlor
Species
rats
rats
rats
rats
rats
rats
Type of
Effect
dvp
dvp
rep
end
rep
rep
Description
NOAEL
LOAEL
NOAEL
NOAEL
NOAEL
LOAEL
Value
50
100
30
50 ~
25
25
Units
mg/kg-day
mg/kg-day
mg/kg
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
S.C., I.V., l.p.,
Inlectlon)
gavage
gavage
gavage
oral gavage
oral
gavage
Exposure
Duration/Timing
gestation days 6-
• 15
gestation days 6-
15
Days 14-20
8 weeks
1 1 months
From gestation,
weaning, lactation
through puberty
Reference
Kheraetal., 1978
Kheraetal., 1978
Qellert and Wilson,
1979
Goldman et.al.,
1986
Gray et al.,jl988
•
Gray et at., 1989
Comments
This NOEL was established by
the authors of the study.
Dose-related increase in wavy
ribs at 1 00, 200, 400 mg/kg-
day.
Reproductive function in the
male or female offspring was
not altered by this dosage.
No effect observed in pituitary
weight, serum LH, FSH, or
prolactin levels and the
pituitary LH of FSH
concentrations.
Reproductive effects were
seen at this level. (See paper
for specific effects on females
and males.)
-------
Freshwater Tox,~..y - Methoxychlor
Cas No. 72-43-5
Chemical
Name
Methoxychlor
Methoxychlor
Methoxychlor
Methoxychlor
Methoxychlor
Species
bluegill
brook trout
rainbow trout
striped bass
fathead
minnow
Type of
Effect
mort
mort
mor
mort
mor
Description
LC50
LC50
LC50
LC50
LC50
Value
40.0 - 75
(50.3)
40
9.36 - 62.6
(33.50)
3.3
7.5 - 35
(11.36)
Units
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NA
NA
NA
NA
NA
Exposure
Duration/
Timing
96 hour
1 day
96 hour
96 hour
96 hour
Reference
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
AQUIRE, 1994
1
Comments
NA = Not applicable
i
-------
Terrestrial Toxicity - Methoxychlqr
Cas No. 72-43-5
Chemical
Name
methoxychlor
methoxychlor
methoxychlor
methoxychlor
,
Species
mouse
rabbit
hamster
duck
Type of
Effect
acute
acute
acute
acute
Description
LD50
LD50
LD50
LD50
Value
1
>6
500
>2
Units.
g/kg
9/kg
mg/kg
g/kg
Exposure
Route (oral,
s.c., l.v.,~l.p.,
Injection)
oral
oral
i.p.
oral
Exposure
Duration/Timing
NS
NS
NS
NS
Reference
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
Comments
NS = Not specified
-------
Terrestrial Toxi. ./ - Methoxychlor
Gas No. 72-43-5
Chemical
Name
methoxychlor
methoxychlor
methoxychlor
methoxychlor
methoxychlor
methoxychlor
methoxychlor
-
• Species
'rats
rats
rabbits
rabbits
mice
cow
rat
Type of
Effect
rep
rep
ter
ter
rep
rep
acute
Description
LOAEL
LOAEL
NOAEL
LOAEL
PEL
NOEL
LD50
Value
100
80
5.01
35.5
200
10
5
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
•
mg/kg-day
9/kg
Exposure
Route (oral,
B.C., I.V., l.p..
Injection)
oral gavage
oral
oral gavage
oral
oral
Exposure
Duration/Timing
70 days (m); 15
days (f)
before mating and
throughout
pregnancy
Days 7 through 1 9
of gestation
Days 7 through 19
ol gestation
5 consecutive days
each week for 4
weeks
NS
NS
Reference
Bal, 1984
Harris etal., 1974
Kincaid .
Enterprises, Inc.
1986 as cited in
IRIS, 1993
Kincaid
Enterprises, Inc.
1 986 as cited in
IRIS, 1993
Martinez and
Swartz, 1992
Macklin and
Ribelin, 1971 as
cited in (ARC, 1979
RTECS, 1994
Comments
The normal reproductive
processes of testes,
epididymis, and ovaries are
impaired at this dosage level.
Impaired reproductive
behavior was observed in
female and male pups whose
mothers were fed this dosage
level (1000 ppm).
Maternal toxiciry was
observed as an excessive loss
of liners (abortions).
The adult ovary appears to be
a target organ for the effects
of MXC. The specific effects
were increased lipid
accumulation in interstitial
cells and theca cells.
No abortions produced in
pregnant cows
-------
APPENDIX B Methyl parathion
Toxicological Profile for Selected Ecological Receptors
Methyl parathion .
Cas No.: 298-00-0
Summary: This profile on methyl parathion summarizes the lexicological benchmarks and
biological uptake measures (i.e., bibconcentration, bioaccumulation, and biomagniflcation factors)
for birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if'available, biomagniflcation factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the freshwater
ecosystem were calculated for organic constituents with log Kow between 4 and 6.5. For the
terrestrial ecosystem, these biological uptake measures also include terrestrial vertebrates and
invertebrates (e.g., earthworms). The entire toxicological data base compiled during this effort
is presented at the end of this profile.. This profile represents the most current information and
may differ from the data presented in the technical support document for the Hazardous Waste
Identification Rule (HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ro) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife which
reported dose-response data. However, toxicological studies involving methyl parathion exposure
to mammals have been conducted using laboratory rats. Lobdell and Johnston (1966 as cited in
NIOSH, 1976) conducted a three-generation rat study involving the dietary administration* of 10
and 30 ppm methyl parathion over a 27 week period. The group of rats fed 30 ppm exhibited
reductions in the survival time of first generation weanlings, decreased number of litters in the
second generation, and elevated total number of stillbirths. However, the authors of this study
concluded that dietary exposure of 30 ppm methyl parathion did not produce a consistent or dose-
related effect on rat reproduction. While it is questionable as to the statistical significance of the
toxic effects at 30 ppm, the observed biological effects are significant enough to warrant a
NOAEL of 10 ppm (equivalent to 1 mg/kg-d in the study). Tanimuria et al. (1967) investigated
August 1995
-------
APPENDIX B Methyl parathion • 2
the embryotoxicity of methyl parathion in rats and mice through the administration of methyl
parathion (suspended in a 0.5% aqueous solution of sodium carboxymethyl) as an single
intraperitoneal injection. Rats were injected once on day 12 of gestation with 5, 10, and 15
mg/kg and mice with 20 and 60 mg/kg methyl parathion on day 10 of gestation. In this study,
embryotoxic effects on rats included suppression of fetal growth and ossification at the 15 mg/kg
dose level (LOAEL). The mice had high fetal mortality and an elevated incidence of cleft palate
as well as suppression of fetal growth at the 60 mg/kg dose level.
The study by Lobdell and Johnston (1966 as cited in NIOSH, 1976) was selected for calculation
of the toxicological benchmark for mammals because it involved chronic exposure over three
generations and it examined effects on a reproductive endpoint. The NOAEL of 10 ppm (1
mg/kg-d) was selected based on the reproductive effects on generations of rats exposed to 30
ppm methyl parathion (Lobdell and Johnston, 1966 as cited in NIOSH, 1976). Other studies
such as Street et al. (1975 as cited in NIOSH, 1976) included more dose levels, but did not
investigate reproductive endpoints, and Tanimuria et al. (1967). used a less preferred
intraperitoneal route of exposure in an acute study with rats and mice.
i
The 1 mg/kg-d dose from Lobdell and Johnston (1966 as cited in NIOSH, 1976) was scaled for
species representative of a freshwater ecosystem using a cross-species scaling algorithm adapted
from Opresko et al. (1994)
Benchmark = NOAEL. x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the default scaling
methodology EPA proposed for carcinogeriicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the study
documented reproductive effects from methyl parathion exposure to both male and female rats,
the mean male and female body weight for each representative species was used in the scaling
algorithm to obtain the toxicological benchmarks.
Data were available on the reproductive and developmental, effects of methyl parathion, as well
as growth or survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations. All of the studies identified were conducted using laboratory
rats and mice or as such, inter-species differences among wildlife species were not identifiable.
Therefore, an inter-species uncertainty factor was not applied. There were no study values in the
data set which were more than a magnitude lower than the benchmark value. Based on the data
set for methyl parathion, the benchmarks were categorized as adequate.
Birds: Several studies were identified concerning reproductive effects observed in birds. Meyers
et al. (1990) exposed red-winged blackbirds to single oral doses of 0, 2.37, 4.21 or 7.5 mg/kg
methyl parathion in propylene glycol during incubation. In this study, no apparent adverse
effects were observed in adult red-wing females at 2.37 or4.21 mg/kg dose levels, but those at
August 1995
-------
APPENDIX B Methyl parathion - 3
the 7.5 mg/kg dose level showed definite signs of intoxication. In another study, Fairbrother et
al. (1988) concluded that methyl parathion administered via gavage at 4 mg/kg (in com oil) of
5-day-old mallard ducks affected the brood-rearing phase of reproduction by direct mortality and
through behavioral changes. Bennett et aJ. (1990) conducted one chronic and one acute test to
examine the effects of dietary methyl parathion exposure on reproduction in bobwhite quail. The
chronic test exposed bobwhite quail to doses of 0, 7, 10, 14, 20, or 28 ppm for a period of 25
weeks. A significant dose-related decrease in the number of eggs laid was observed for
concentrations greater than 10 ppm. Also at concentrations greater 10 ppm, there was a dose-
related decrease in eggshell weight per unit area. The ppm dose in the Bennett et al. (1990)
study was converted to a daily dose using a food ingestion rate calculated using an allometric
equation based on a body weight of the test species (Nagy, 1987). Assuming a body weight of
0.180 kg (Roseberry and Klimistra, 1971), doses from the Bennett et al. (1990) study were
calculated as 0.74, 1.05, 1.5, 2.1, or 3.0 mg/kg-day, with a NOAEL of 1.05 mg/kg-day for
reproductive effects. In a later study, Bennett et al. (1991) exposed seven-month-old mallards
to a dietary dose of 400 ppm for eight days to evaluate egg laying and, incubation during the
nesting cycle. The results of this study demonstrated that the nesting success my be impacted
by short dietary exposures to methyl parathion during early incubation. Buerger et al. (1991)
studied the effects of 2, 4, and 6 mg/kg methyl parathion on Northern bobwhite survivability for
three field seasons. Bobwhites receiving oral dose of 6 mg/kg methyl parathion had lower
survival than control birds due to predation, not overt toxicity. Using a bobwhite quail body
weight of 0.180 kg (Roseberry and Klimistra, 1971) and the food ingestion equation from Nagy
(Nagy, 1987), the 6 ppm level was converted to a daily dose of 0.42 mg/kg.
r-
The study by Bennett et al. (1990) was selected for extrapolation of a benchmark because: (1)
exposure occurred a critical time in the reproductive cycle, (2) the dose range was sufficient to
establish a dose-response curve, and (3) the study evaluated a reproductive endpoint. Although
the Buerger et al. (1991) study derived a slightly lower value than the benchmark, it was not
chosen as the benchmark because the results from two years did not follow a normal dose-
response relationship and the decreased survival rate at 6 mg/kg was primarily due to increased
predation, rather than overt toxicity effects of methyl parathion. The Bennett et al. (1991) and
the Fairbrother et al. (1988) studies were not selected because they did not establish a dose-
response relationship, although each study did examine a reproductive endpoint. The NOAEL
from the Meyer et al. (1990) study was not selected because it was not the lowest NOAEL in the
data set.
The chronic dose from the Bennett et al. (1990) study was then scaled for species representative
of a freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994). This is the default scaling methodology EPA proposed for carcinogenicity assessments
and reportable quantity documents for adjusting animal data to an equivalent human dose (57 FR
24152). Since the study documented reproductive effects from methyl parathion exposure on
both male and female quail, body weights for male and female representative species was used
in the scaling algorithm to obtain the toxicological benchmarks.
Data were available on reproductive and developmental effects of methyl parathion, as well as,
on growth or survival endpoints. In addition, the data set contained studies which were
August 1995
-------
APPENDIX B Methyl parathion - 4
conducted over chronic and subchronic durations and during sensitive life stages. Laboratory
experiments of similar types were not conducted on a wide range of avian species and as such,
inter-species differences among wildlife species were not identifiable. There were no values in
the data set which were more than an order of magnitude lower than the benchmark value.
Based on the avian data set for methyl parathion, the benchmarks that were developed were
categorized as adequate.
Fish and aquatic invertebrates: Since a Final Chronic Value (FCV) did not exist for methyl
parathion, a Secondary Chronic Value (SCV) of 3.2E-5 mg/1 was calculated using the Tier II
methods described in Section 4.3.5. Because the benchmark for daphnids was calculated using
the Tier II method, the benchmark was categorized as interim.
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECM) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum). Aquatic
plant data was not identified for methyl parathion and, therefore, no benchmark was developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbon partition coefficient (K^) to determine a protective sediment concentration (Stephan,
1993). The EQp number is the chemical concentration that may be present in sediment while still
protecting the benthic community from the harmful effects of chemical exposure. Since a FCV
for methyl parathion was not available, a Secondary Chronic Value (SCV) was calculated as
described in Section 4.3.5. The SCV was used to calculate an EQp number of 2.1E-2 mg methyl
parathion /kg organic carbon. Assuming a mass fraction of organic carbon for the sediment (f^.)
of 0.05, the benchmark for the benthic community is 1.04E-3 mg/kg. Since the EQp number was
based on a SCV, and not an FCV, the sediment benchmark is categorized as interim.
August 1995
-------
APPENDIX B
Methyl parathion • 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
R*j>ra**flta13v«
SpacJW
mink
otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sand piper
herring gull
kingfisher
Soocnmark
Value* mB/Ka-
day
0.80 (a)
0.48 (a).
0.49 (a)
0.61 (a)
0.55 (a)
0.66 (a)
0.73 (a)'
1.5 (a)
0.67 (a)
1.10(a)
StiKJy
Spsciw
rat
rat
bobwhite quail
bob white quail
bobwhite quail
bobwhite quail
bobwhite quail
bobwhite quail
bobwhite quail
bobwhite quail
Wect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study
VaJuo
mg/kg-day
1
1
1.05
1.05
. 1.05.
'l.05
1.05
1.05
1.05
1.05
; Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
$F
-
-
-
-
•
Prtoln*! Souro*
Lobdelland
Johnston, 1966 as
cited in NIOSH,
1976
Lobdeland '
Johnston, 1966 as
cited in NIOSH,
1976
Bennett et al.,
1991
Bennett et al.,
1991
Bennett et al.,
1991
Bennett et al.,
1991
Bennett et al.,
1991
Bennett et al.,
1991
Bennett et al.,
1991
Bennett et al.,
1991
•Benchmark Category, a = adequate, p • provisional, i
above the NEL or LEL for other adverse effects.
interim; a '" indicates that the benchmark value was an order of magnitude or more
August 1995
-------
APPENDIX B
Methyl parathion - 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ftopr««»ntatfv«
Specie*
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark
VahiV
rag/L
3.2 E-05 (i)
ID
3.9 E-04 (i_
mg/kg sediment
Study SpaciM
. AWQC
Species
-
benthic community
Description
scv
SCV x K^
Original '
Souro*
. AQUIRE, 1995
-
AQUIRE. 1995
II.
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rational behind Toxicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As discussed previously in the freshwater ecosystem discussion, no suitable subchronic
or chronic studies were found for mammalian wildlife exposure to methyl parathion. Since no
additional studies for terrestrial mammals were found, the same surrogate-species study (Lobdell
and Johnston, 1966 as cited NIOSH, 1976) was used to calculate benchmark values for
mammalian species representative of terrestrial ecosystems. This value was then scaled for
species representative of a terrestrial ecosystem using a cross-species scaling algorithm adapted
from Opresko et al. (1994). Based on the data set for methyl parathion, the benchmarks
developed from Lobdell and Johnston (1966 as cited NIOSH, 1976) were categorized as
adequate.
Birds: No additional avian toxicity studies were identified for species representative of the
terrestrial ecosystem. Therefore, the study by Bennett et al. (1990) was selected for the
extrapolation of a benchmark for the representative bird species of a terrestrial environment. The
NOAEL of 1.05 mg/kg-day was scaled for species representative of a terrestrial ecosystem using
a cross-species scaling algorithm adapted from Opresko et al. (1994). Since the study
documented reproductive effects from methyl parathion exposure on both male and female
bobwhite quail, body weights for male and female representative species was used in the scaling
August 1995
-------
APPENDIX B Methyl parathion - 7
algorithm to obtain the lexicological benchmarks. Based on the avian data set for methyl
parathion, the benchmarks that were developed were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks,
were selected by rank ordering the LOEC values and then approximating the 10th percentile. If
there were 10 or fewer values for a chemical, the lowest LOEC was used. If there were more
than 10 values, the 10th percentile LOEC was used. Such LOECs applied to reductions in plant
growth, yield reductions, or other effects reasonably assumed to impair the ability of a plant
population to sustain itself, such as a reduction in seed elongation. However, terrestrial plant
studies were not identified for methyl parathion and, as a result, a benchmark could not be
developed.
Soil community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Methyl parathion - 8
Table 3. Toxicological Benchmarks .for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
BeprMvattMhr*
Specie*
deer mouse
short-tailed .shrew
meadow vole
Eastern cottontail
red fox
raccoon
white tailed deer
red-tailed hawk
American kestrel
Northern bobwhite
American robin
American woodcock
plants
soil community
Benchmark
Value* njg/kfl-
tf«y
2.2 (a)
2.2 (a) .
1.9 (a)
0.76 (a)
0.55 (a)
0.52 (a)
0.26 (a)
0.66 (a)
1.16 (a)
1.06 (a)
1.28 (a)
1.07 (a)
ID
ID
Study
Sped**
rat
rat
rat
rat
rat
rat
rat
bobwhite quail
bobwhite quail
3obwhite quail
bobwhite quail
>obwhite quail
•
*
elect
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
-
Study
Value
rnfl/fcg-day
1
1
1
1
1
1
1
1.05
1.05
1.05
1.05
1.05
•
Description
NOAEL
NOAEL
NOAEL
.NOAEL .
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
•
"
SF
•
•
*
•
*
*
OHfllrwl $awe* \
Lobdell and Johnston,
1966 as cited in
NIOSH, 1976
Lobdell and Johnston,
1966 as cited in
NIOSH, 1976
Lobdell and Johnston,
1966 as cited in
NIOSH, 1976
Lobdell and Johnston,
1966 as cited in
NIOSH, 1976
Lobdell and Johnston,
1966 as citad in
NIOSH, 1976
Lobdell and Johnston,
1966 as cited in
NIOSH, 1976
Lobdelt and Johnston,
1966 as cited in
NIOSH, 1976
Bennett et al.. 1990
Bennett eta!., 1990
Bennett etal., 1990
Bennett et al.. 1990
Bennett etal., 1990
-
'Benchmark Category, a = adequate, p = provisional, i = interim; a '" indicates that the benchmark value was an order of magnitude
or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Methyl parathion - 9
in. Biological Uptake Measures
This section presents biological uptake measures (i.e, BCFs, BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for selected ecological receptor categories: fish in the limnetic
.or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates, terrestrial
vertebrates, and plants. For the generic aquatic ecosystems, the BCF value is identified as whole-
body or lipid-based and designated with a "d" if the value reflects dissolved water concentrations,
and a "t" if the value reflects total surface water concentrations. For organic chemicals with log
KOW values below 4, bioconcentration factors (BCFs) in fish were always assumed to refer to
dissolved water concentrations (i.e., dissolved water concentration equals total water
concentration). The following discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values presented in
Table 4.
The bioconcentration factor for fish was estimated from the Thomann (1989) model (i.e., log Kow
~ dissolved BCF/) because: (1) only two measured values were available, (2) the predicted BCF
was within a factor of 2 of the geometric mean of measured BCFs (i.e., the difference was
insignificant), (3) the BCF was in close agreement with predicted BCFs based on other methods
(i.e., regression equations), and (4) there were no data (e.g., metabolism) to suggest that the log
Kow = BCF;d relationship deviates for methyl parathion (log Kow = 2.86). As stated.in section
5.3.2, the dissolved bioconcentration factor (BCF,d ) for organic chemicals with log Kow below
4 was considered to be equivalent to the total bioconcentration factor (BCF/) and, therefore,
adjusting the BCFjd by the dissolved fraction (fd) was not necessary.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation method
is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is dominated
by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial wildlife developed
for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial Wildlife from TCDD and
TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient quality to serve as the standard.
The beef biotransfer factor (BBFs) for a chemical lacking measured data is compared to the BBF
for TCDD and that ratio (i.e., methyl parathion BBF/TCDD BBF) is multiplied by the TCDD
standard for terrestrial vertebrates, invertebrates, and earthworms, respectively. For hydrophobic
organic constituents, the bioconcentration factor for plants was estimated as described in Section
6.6.1 for above
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
-------
ARPENDIX B
Methyl parathion - 10
Table 4. Biological Uptake Properties
•cotogjcai
receptor
fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
SCF, BAF, or
BSAF
BCF
BAF
BAF
BCF
BCF
BCF
irpld-baaed of
wftote-body
lipid
lipid
whole-body
whole-body
whole-body
whole-plant
value
723 (d or t)
8.9E-06
8.6E-06
6.8E-05
0.86
•oure*
predicted value based on
Thomann, 1989
insufficient data
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
U.S. EPA. 1992e
d » refers to dissolved surface water concentration
t a refers to total surface water concentration
August 1995
-------
APPENDIX B Methyl parathion - 11
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
AQUIRE (AQUntie Toxicity /nformation /?£trieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
. Duluth, MN,
Bennett, R.S., B.A. Williams, D.W. Schmedding, and J.K. Bennett. 1991. Effects of Dietary
Exposure to Methyl Parathion on Egg Laying and Incubation in Mallards. Environ.
Toxicol. Chem., 10:501-507.
Bennett, R.S., R. Bentley, T. Shiroyama, and J.K. Bennett. 1990. Effects of the Duration and
Timing of Dietary Methyl Parathion Exposure on Bobwhite Reproduction. Environ.
Toxicol. Chem., 9:1473-1480.
Brewer, L.W., C.J. Driver, R.J. Kendall, C. Zenier, and T.E. Lacher, Jr. 1988. Effects of
Methyl Parathion in Ducks and Duck Broods. Environmental Toxicology and Chemistry,
Vol. 7, pp. 375-379.
Buerger, T.T., R.J. Kendall, B.S. Mueller, T. DeVos, and B.A. Williams. 1991. Effects of
Methyl Parathion on Northern Bobwhite Survivability. Environmental Toxicology and
Chemistry, Vol. 10, pp. 527-532.
Carter, F.L., and J.B. Graves. 1972. Measuring Effects of Insecticides on Aquatic Animals.
La. Agric. 16(2): 14-15. As cited in AQUIRE (AOUatic Toxicity Information REtrieval
Database). Environmental Research Laboratory, Office of Research and Development,
U.S. Environmental Protection Agency. Duluth, MN.
Grassland, N.O., and D. Bennett. 1984. Fate and Biological Effects of Methyl Parathion in
Outdoor Ponds and Laboratory Aquaria. I. Ecotoxicol. Environ. Saf. 8(5):471-481. As
cited in AQUIRE (AQUatic Toxicity Information REtrieval Database). Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency. Duluth, MN.
De Bruijn, J., and J. Hermans. 1991. Uptake and Elimination Kinetics of Organophosphorus
Pesticides in the Guppy (Poecilia reticulata): Correlations with the Octanbl/Water
Partition. Environ. Toxicol. Chem. 10(6):791-804. As cited in AQUIRE (AOUatic
Toxicity Information REtrieval Database). Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency. Duluth, MN.
August 1995
-------
APPENDIX B Methyl parathion - 12
Degraeve, N., M.C. Chollet, and J. Moutschen. 1984. Cytogenetic Effects induced by
Organophosphorus Pesticides in Mouse Spermatocytes. Toxicology Letters, 21:315-319.
Delnicki, D., and K.J. Reinecke. 1986. Mid-winter food use and body weights of mallards
and wood ducks in Mississippi. J. Wildl. Manage. 50:43-51.
Dortland, R.J. 1980. Toxicological Evaluation of Parathion and Azinphosmethyl in Freshwater
Model Ecosystems. Versl. Landbouwkd. Onderz 898:1-112. As cited in AQUIRE
(AQUatic Tpxicity Information REtrieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency.
Duluth, MN.
Fairbrother, A., S.M. Meyers, and R.S. Bennett. 1988. Changes in Mallard Hen and Brood
Behaviors in Response to Methyl Parathion-Induced Illness of Ducklings. Environ.
Toxicol.Chem.,l:499-5Q3. '
Hansch, C., and A.J. Leo. 1985. Medchem Project, Issue No. 26. Claremont, CA: Pomona
College. As cited in Howard, P.H. .1991. Handbook of Environmental Fate and
Exposure Data for Organic Chemicals. Volume III: Pesticides. Lewis Publishers.
Chelsea, Michigan.
Henry, M.G. and G.J. Atchison. 1984. Behavorial Effects of Methyl Parathion on Social
Groups of Bluegill (Lepomis macrochirus). Environmental Toxicology and Chemistry,
Vol. 3, pp. 399-408.
Jarvinen, A.W., and D.K. Tanner. 1982. Toxicity of Selected Controlled Release and
Corresponding Unformulated Technical Grade Pesticides to the Fathead Minnow
Pimephalas promelas. Environ. Pollut. Ser. A Ecol. Biol. 27(3): 179-195. As cited in
AQUIRE (AQUatic Toxicity Information REtrieval Database). Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency.
Duluth, MN. • .
Johnson, W.W. and M.T. Finley. 1980. Handbook of Acute Toxicity of Chemicals to Fish
and Aquatic Invertebrates. Resour. Publ. 137, Fish Wildl. Serv., U.S.D.I., Washington,
DC, p. 98. As cited in AQUIRE (AQUatic Toxicity Information REtrieval Database).
Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency. Duluth, MN.
Lobdel, B.J. and C.D. Johnston. 1966. Methyl Parathion - Three Generation Reproduction
Study in the Rat. Woodard Research Corp., Hemdon, VA. As cited in NIOSH (National
Institute for Occupational Safety and Health). 1976. Criteria for a Recommended
Standard for Occupational Exposure to Methyl Parathion. U.S. Department of Health,
Education and Welfare, Washington, DC.
August 1995
-------
APPENDIX B Methyl parathion - 13
Metcalf, R.L. et al. 1979. Design and Evaluation of a Terrestrial Model Ecosystem for
Evaluation of Substitute Pesticide Chemicals, pp. 308. U.S. EPA-600/3-79-004. As cited
in Howard, P.H. 1991. Handbook of Environmental Fate and Exposure Data for Organic
Chemicals. Volume IE: Pesticides. Lewis Publishers. Chelsea, Michigan.
Meyers, S.M., J.L. Cummings, and R.S. Bennett. 1990. Effects of Methyl Parathion Red-
Winged Blackbird (Agelaiusphoeniceus) Incubation Behavior and Nesting Success.
Environmental Toxicology and Chemistry, 9:807-813.
Nagy, K.A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol.Mono. 57:111-128.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. lexicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Palawski, D., D.R. Buckler, and F.L. Mayer. 1983. Survival and Condition of Rainbow
Trout (Salmo gairdneri) After Acute Exposures to Methyl Parathion, Triphenyl Phosphate
and DEF. Bull. Environ. Contain. Toxicol. 30(5):614-620. As cited in AQUIRE
(AQUatic Toxicity information REtrieval Database). Environmental Research Laboratory,
Office of Research and Development, U.S. Environmental Protection Agency. Duluth,
MN.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Robinson, S.C., R.J. Kendall, R. Robinson. 1988. Effects of Agricultural Spraying of Methyl
Parathion on Cholinesterase Activity and Reproductive Success in Wild Starlings.
Environmental Toxicology and Chemistry, Vol. 7, pp. 343-349.
Roseberry and Klimistra. 1971. Annual weight cycles in male and female bobwhite quail.
Aw* 88:116-123.
Stephan, C.E. 1993. Derivations of proposed human health and wildlife bioaccumulation
factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter n, G.W., M.A. Futrell, and G.A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management. U.D. Department of Energy,. Washington, DC.
August 1995
-------
APPENDIX B Methyl parathion - 14
Suter n, G.W. and J.B. Mabrey. 1994. Toxicological benhmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-96/R1.
U.S. Department of Energy, Oak Ridge National Laboratory, Oak Ridge, TN.
Tanimura, T., T. Katsuya, and H. Nishimura. 1967. Embryotoxicity of Acute Exposure to
Methyl Parathion in Rats and Mice. Arch. Environ. Health, Vol. 15.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology and Chemistry 11:615-629.
U.S.EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. EPA/600/6-87/008.
Environmental Criteria and Assessment Office, Office of Health and Environmental
Assessment, Office of Research and Development, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment. Washington, D.C. January.
U.S. EPA (Environmental Protection Agency). 1993. Technical Basis for Deriving Sediment
Quality Criteria for Nonionic Organic Contaminants for the Protection of Benthic
Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of Water,
Washington, DC.
Vamagy, L., R. Imre, T. Fancsi, A. Hadhazy. 1982. Teratogenicity of Methyl Parathion 18
WP and Wofatox 50 EC in Japanese Quail and Pheasant Embryos, with Particular
Reference to Osteal and Muscular Systems. Acta Veterinaria Academiae Scientiarum
Hungaricae, Vol. 30 (1-3), pp. 135-146.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U:S. Department of Energy.
August 1995
-------
Terrestrial Toxic. Methyl parathfon
Cos No.: 298-00-0
Chemical
Name
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Species
mallard
mallard
mallard
mallard
mallard
^
*
Endpoint
behv
behv, rep
behv, mort.
behv
behv
behv, rep
Description
NOAEL
AEL
AEL
NOAEL
NOAEL
NOAEL
Value
300
400
4
300
300
1.05
Units
ppm
ppm
mg/kg-
bodywt.
ppm
ppm
mq/kq-dav
Exposure Route
(oral, s.c., i.v., i.p ,
injection)
oral
oral
oral (gavage)
oral
oral
oral
Exposure Duration
/ Timing
8 days
'egg (aying": 8
days initiated after
the 4th egg laid
one dose at 5-6
days old
'early incubation':
8 days initiated
after 4th day of
incubation
'late incubation': 8
days initiated after
16th day of
incubation
25 wks (10 wks
prior to egg-laying
6 wks coming into
laying, 9 wks
during laying)
Reference
-
Bennett etal., 1991
*
Bennett etal., 1991
Fairbrother et al., 1988
Bennett etal.. 1991
Bennett etal., 1991
Bennett etal., 1990
Comments
Incubation behv / Dose-resp is based on 2
LC50 values, the 300ppm and £ 400 ppm (all
dietary); food cons (controls) 115 g/bird-day
Nest abandonment, incr mortality (hens), deci
egg laying per hen / Insufficient dose-resp
info- only 2 values tested (300 & 400); food
cons (controls) 115g/bird-day
Methyl parathion affected the brood-rearing
phase of reproduction by direct mortality and
through behavorial changes.
Nest abandonment, incr mortality (hens), deci
diet / Insufficient dose-resp'info- only 2
values tested (300 & 400); food cons
(controls) 115 g/bird-day
Mortality (hen), nest abandonment, deer diet /
Insufficient dose-resp info- only 2 values
tested (300 & 400); food cons (controls) 115
g/bird-day
Deer diet (84% of controls), deer egg laying,
deer eggshell weight per unit area, deer brain
ChE / dose-resp established; food cons
(controls) 21.7 q/bird-day
-------
Terrestrial Toxicity - Methyl parathlon
Cos No.: 298-00-0
Chemical
Name
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Paiathion
Species
bobwhite
red-winged
blackbird
red-winged
blackbird
rat
mice
rat
rat
Endpoint
behv, rep
behv-
behv
dev
dev
rep.
mort.
Description
LOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LD50
Value
1.06
2.37
4.21
15
60
1
6010
Units
mg/kg-day
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg-day
ug/kg-
body wl.
Exposure Route
(oral, s.c., i.v.. i.p.,
injection)
oral
oral
oral
i.p.
i.p.
oral (in food)
oral
Exposure Duration
/ Timing
6 wks (3 wks
during egg laying,
3 wks following
laying)
single dose
single dose
one injection on
day 12 of
gestation
one injection on
day 10 of
gestation
27 weeks
NS
Reference
Bennett et al., 1990
Meyers et al., 1990
Meyers etal., 1990
Tanimura etal., 1967
Tanimura et al., 1967
Lobdell and Johnston, 1966
as cited in NIOSH, 1976
RTECS, 1994
Comments
Deer diet (63% of controls), deer egg laying,
deer eggshell weight per unit area, strength &
thickness / dose- response questionable (no
NOAEL value identified, unlike above longer
term study); food cons (controls) 21.7 g/bird-
day
Brain ChE depression, time away from nest /
does-resp, 2.37 and 4.21 only values tested
Brain ChE depression, time away from nest:
2.25hr versus 1 . 1 5 for 2.37-group (possible
reduction in fledgling survival if predators
exist) / does-resp questionable 2.37 and 4.21
only values tested
Embryotoxic effects included suppression of
fetal growth and ossification after the
admisitration of (tie adult dose
Embryotoxic effects included incidence of
cleft palate. suppression of fetal growth, and
ossification after the admisiiration of the adult
dose
99% pure methyl parathion was administered
at doses of 3 mg/kg-d and 1 mg/kg-d for this
3-generation study. At 1 mg/kg-d (her was no
consistent or dose-related effects on
reproduction, however there was reduced
survival of F3 weanlings.
-------
Freshwater Biological Upto. Measures - Methyl parathion
Cos No.: 298-00-0
Chemical Name
Methyl parathion
Methyl parathion
Species
rainbow trout
quppv
B-factor
(BCF, BAF,
BMP)
BCF
BCF
•
Value
12-71
0.0959
Measured or
predicted
(m,p)
m
m
Units
NS
NS
Reference
Grassland and Bennett,
1984 as cited in
AQUIRE, 1994
DeBruijn and Hermens,
1991 as cited in
AQUIRE, 1994
Comments
14 day test
6-7 mo. 40-280 MG; 3 to 1 1
day test
-------
Terrestrial Biological Uptake Measures - Methyl parathion
CasNo.: 298-00-0
Chemical Name
Methyl parathion
Species
plants
B-factor
(BCF.BAF,
BMP)
BCF
Value
5.6
Measured or
Predicted (m.p)
p
Units
(ug/g WW
plant)/(ug/mL soil
water)
Reference
U.S. EPA. 1990e
Comments
Plant uptake from
soil pertains to leafy
vegetables
-------
Terrestrial Toxic. Methyl parathion
CasNo.: 298-00-0
Chemical
Name
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Methyl
Parathion
Species
dog
rabbit
guinea pig
duck
mammal
wild bird
Endpoint
mort.
mort.
mort.
mort.
mort.
mort.
Description
LD50
LD50
LD50
LD50
LD50
LD50
Value
90
420
1270
10
57
5
Units
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
body wt.
mg/kg-
bodywt.
Exposure Route
(oral, s.c., i.v., i.p.,
injection)
oral
oral
oral
oral
oral
oral
Exposure Duration
/ Timing
NS
NS
NS
NS
NS
NS
Reference
RTECS, 1994
RTECS, 1994
-
RTECS. 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
Comments
-------
Freshwater Toxiclty - Methyl parathfon
Cas. No.: 298-00-0
Chemical
Name
Methyl
parathion
Methyl
parathion
Methyl
parathion
Methyl
parathion
Methyl
parathion
Species
channel
catfish
bluegill
rainbow trout
daphnia
magna
fathead
minnow
Endpolnt
mo rt.
mod.
mort.
immob.
mort.
Description
LC50
LC50
LC50
EC50
LC50
Value
5240
1600
2800
(7.8-9.1)
8.34
8900
Units
ug/L
ug/L
ug/L
ug/1
uq/L
Test Type
(static/ (low
through)
NS
NS
NS
NS
NS
Exposure
Duration/
Timing
4 days
4 days
4 days
48 hours
4 days
Reference
Johnson and Finley,
1980 as cited in
AQUIRE. 1995
Carter and Graves,
1972 as cited in
AQUIRE, 1995
Palawski et al , 1983
as cited in AQUIRE,
1995
Portland, 1980 as
cited in AQUIRE,
1995
Jarvinen and Tanner,
1982 as cited in
AQUIRE, 1995
Comments
t
-------
APPENDIX B Molybdenum - 1
lexicological Profile for Selected Ecological Receptors
Molybdenum
Cas No.: 7439-98-7
Summary: This profile on molybdenum summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors)
for birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the freshwater
ecosystem were calculated for organic constituents with log Kow between 4 and 6.5. For the
terrestrial ecosystem, these biological uptake measures also include terrestrial vertebrates and
invertebrates (e.g., earthworms). The entire toxicological data base compiled during this effort
is presented at the end of this profile. This profile represents the most current information and
may differ from the data presented in the technical support document for the Hazardous Waste
Identification Rule (HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C_) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
' .
Study Selection and Calculation of Toxicological Benchmarks '
Mammals: Several studies were identified which investigated molydenum-induced toxicity in
mammalian species. In a multi-generational study, mice were exposed orally to 1.18 mg Mo/kg-
day (Schroeder and Mitchener, 1971). Reproductive and fetotoxic effects exhibited by the third
generation included increased infertility in the mating pairs and excess fetal mortality among the
offspring. Since only one dose was used for this study, an AEL of 1.18 mg/kg-day was inferred.
In a two-part study, rats and rabbits were exposed to oral daily doses of molybdenum ranging
from 500 to 2000 ppm (Arrington et al., 1965). Although rats exposed for six-weeks to 500 ppm
showed no signs of clinical toxicity, those given 1000 ppm had reduced voluntary feed intake,
decreases in growth and feed utilization efficiency. Based on these results, a NOAEL of 500
ppm and a LOAEL of 1000 ppm were inferred for molybdenum toxicity in rats. Rabbits-exposed
August 1995
-------
APPENDIX B Molybdenum - 2
for three weeks to 2000 ppm exhibited similar signs of toxicity including reduced voluntary feed
intake and growth while those rabbits given 1000 ppm showed no adverse effects. A NOAEL
of 1000 ppm and a LOAEL of 2000 ppm for pathological effects of molybdenum in rabbits.
Fungwe et al. (1990) exposed rats to molybdenum in drinking water at doses of 5, 10, 50 or 100
mg/L. In addition, all groups of rats were fed a diet containing 0.025 mg/kg of molybdenum
inherent to the diet. The exposure period extended from six weeks prior to mating through day
21 of gestation. No signs of toxicity were observed in rats given 5 mg/L however, .those given
10 mg/L exhibited an increasing incidence of resorbed fetuses and sites of resorption and a
decrease in average litter size. A NOAEL of 5 mg/L and a LOAEL of 10 mg/L can be inferred
for fetotoxic effects. The NOAEL inferred from the Fungwe et al. (1990) study needed to be
converted from mg/kg and mg/L to mg/kg-day. The following equation was used to convert
exposure from the molybdenum inherent to the diet in mg/kg to mg/kg-day:
Food Consumption = 0.056(W°'6611) where W is body weight in kg (U.S. EPA, 1988).
Assuming an average weight of 0.020 kg (U.S. EPA, 1988), the exposure dose from the diet was
estimated to be 0.005 mg/kg-day. The exposure to molybdenum from drinking water was
calculated by first determining the geomean of the molybdenum intake from water intake,
0.72mg/l per week, which was then converted to a daily dose of 0.103 mg/kg-day. Adding the
two exposures together did not change the dose levels significantly, as 0.005 is only 5% of the
exposure from drinking water. Therefore, the NOAEL of 5 mg/L was estimated to be equivalent
to 1.03 E-01 mg/kg-day.
Although the Schroeder and Mitchener (1971) study investigates reproductive effects of
molybdenum exposure in mice, it was not considered suitable for the derivation of a benchmark
value as multiple levels of exposure were not utilized and, therefore, a dose-response relationship
was not established. The Arlington et al. (1965) study does provide a dose-response relationship
for molybdenum toxicity in rats and rabbits however, the lexicological endpoints do not clearly
indicate that a wildlife population's fecundity would be impaired.
The study by Fungwe et al. (1990) is considered the most suitable for derivation of a mammalian
lexicological benchmark since (1) a dose-response relationship is established, and (2) the study
focuses on reproductive or fetotoxic endpoints. This value was then scaled for species
representative of a freshwater ecosystem using a cross-species scaling algorithm adapted from
Opresko et al. (1994):
( bw V4
Benchmarkw = NOAEL. x
where NOAELt is the NOAEL (or LOAEL/ 10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Fungwe et al.
(1990) study documented fetotoxic effects of molybdenum on female rats, female body weights
August 1995
-------
APPENDIX B Molybdenum - 3
for each representative species were used in the scaling algorithm to obtain the lexicological
benchmarks. Based on the data set for molybdenum, the benchmarks developed from the Fungwe
et ah, (1990) study were categorized as adequate.
Birds: Data involving molybdenum toxicity in avian species were not identified and therefore
benchmarks for avian species could not be derived.
Fish and aquatic invertebrates: Since an AWQC was not available for molybdenum, the Tier
II methodology described in Section 4.3.5 was used to calculate a Secondary Chronic Value
(SCV). Suter and Mabrey (1994) reported an SCV of 2.4 E-01 mg/1. tier II values are
developed so that aquatic benchmarks can be derived for chemicals lacking the necessary data
to calculate an FCV. The SCV of 2.4 E-01 mg/1 was selected as the benchmark protective of
daphnids, fish, and other aquatic organisms. Because the benchmark is based on an SCV, rather
than an FCV, the value was categorized as interim.
Aquatic Plants: The, benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or 2) an effective concentration (EC,,) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastnun capricornutwri). No CV was reported for
molybdenum and, therefore, no benchmark was developed. As described in Section 4.3.6, all
benchmarks for aquatic plants were designated as interim.
Benthic community: The molybdenum benchmark protective of benthic organisms is pending
a U.S. EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Molybdenum • 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with a Freshwater Ecosystem
R«fKM*aMiv»
Sptetw
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Btfluhumh
VMiftA tnttltftt^t
• 0.08 (a)
0.04 (a)
ID
ID
ID
ID
ID
ID
ID
ID
Study
^ SpntM
rat
rat
•
-
-
-
-
£ftet
fat
fet
-
-
-
-
•
SladyVrt*
m0*04i
0.10
0.10
-
-
-
-
-
DMCrtptian
NOAEL
NOAEL
-
-
•
-•
8F
-
-
-
•
-
-
-
Original Source
Fungwe et al.,
1990
Fungwe et al.,
1990
-
-
-
-
-
-
•Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (*) indicates that the benchmark value was an
order of magnitude or more above the NEL or LEL for other adverse effects.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
xsr
fish and aquatic
invert erbrates
aquatic plants
benthic community
-%***
2.4 E-01 (i)
-
under review
Study
aquatic
organisms
aquatic
plants
-
Vtfu* i
2.4 E-01
~*
scv
-
Suter & Mabrey,
1994
•Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data: a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Molybdenum - 5
II. lexicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial ecosystem.
Study Selection and Calculation of lexicological Benchmarks
Mammals: Since no additional mammalian toxicity studies were identified, the Fungwe et al.
(1990) study used to calculate a benchmark value for mammalian species in the freshwater
ecosystem was also used to calculate a mammalian benchmark value for species in the terrestrial
ecosystem. As with the freshwater benchmark calculation, the study NOAEL of 1,03 E-01
mg/kg-day was scaled for species in the terrestrial ecosystem using the cross-species scaling
algorithm adapted from Opresko et al. (1994). Since the Fungwe et al. (1990) study documented
fetotoxic effects from molybdenum exposure to female rats, female body weights for each
representative species were used in the scaling algorithm to obtain the lexicological benchmarks.
Based on the data set for molybdenum, the benchmarks developed from the Fungwe et al. (1990)
study were categorized as adequate.
Birds: As mentioned in the freshwater ecosystem discussion, data involving molybdenum toxicity
in avian species were not identified.
Plants: Molybdenum is essential to plant growth and development, but there is a narrow range
between its concentration as a nutrient and a toxicant Adverse effects levels for terrestrial plants
were identified for endpoints ranging from percent yield to root length. As presented in Will and
Suter (1994), phytotoxicity benchmarks were selected by rank ordering the LOEC values and then
approximating the 10th percentile. If there were 10 or fewer values for a chemical, the lowest
LOEC was used. Such LOECs applied to reductions in plant growth, yield reductions, or other
effects resonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. However, terrestrial plant studies were not identified for
molybdenum, and as a result, a benchmark could not be developed.
i
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Molybdenum - 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
RaprwMnlatfv*
SpMfe*
deer mouse
short-tailed
shrew
meadow vole
Eastern
Cottontail
red fox
raccoon
white- tailed
deer
red-tailed
hawk
American
kestrel
Northern
bobwhite
American
robin
American
woodcock
plant
soil community
*it-,.-ti— i-.t-
tMRIOfwIWK
Vriu* ntoifca-*
0.19 (a)
0.1 9 (a)
0.15 (a)
0.07 (a)
0.05 (a)
0.05 (a)
0.02 (a)
ID
ID
ID
ID
ID
ID
ID
Study
SMfthtt
rat
rat
rat
rat
rat
rat
, rat
-
ElMDt
fet
fet
fet
fet
fet
fet
fet
-
-
- '•
-
•
-
•
Study
Virtu*
*gflf*4
0.10
0.10
0.10
0.10
0.10
0.10
0.10
-
:
OMCdpUOft
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
-
•
8F
-
-
-
-
•
-
-
-
OrigltMidowc*
Fungwe et al.,
1990
Fungwe et al.,
1990
Fungwe et al.,
1990
Fungwe et al.,
1990
Fungwe et al.,
1990
Fungwe et al.,
1990
Fungwe et al.,
1990
•
-
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = interim, ID = insufficient data: a '" indicates that the benchmark value was an
order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Molybdenum - 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for ecological receptor categories: fish in the limnetic or littoral
ecosystem, aquatic invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and
plants. For metals, BCFs are whole-body bioconcentration factors and refer to total surface water
concentrations (versus freely dissolved concentrations). Consequently, all calculations of
acceptable tissue concentrations (TC) represent whole-body concentrations. The following
discussion describes the rationale for selecting the biological uptake factors and provides the
context for interpreting the biological uptake values.
Insufficient data were identified to determine a BCF value in fish, aquatic invertebrates,
terrestrial vertebrates, terrestrial invertebrates and earthworms. A whole plant BCF value of 8.5
E-01, was derived from U.S. EPA (1992e). For metals, empirical data, were used to derive the
BCF for aboveground forage grasses and leafy vegetables. In particular, the uptake response
slope for forage grasses was used as the BCF for plants in the terrestrial ecosystem since most
of the representative plant-eating species feed on wild grasses.
Table 4. Biological Uptake Properties
^4H*l4M«b^«i
•OwVO^pRW
fMMptOT
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
8CF,8AF,ar
98&
•
•
•
•
-
BCF
%M*MMior
VMthjdAliuhdfe*
•fiivivvvoy
-
-
•
-
-
whole-plant
*•*»
ID
ID
ID
ID
ID
8.5 E-01
*oerc*
-
.
-
-
U.S. EPA. 1992e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
ID = refers to insufficient data
August 1995
-------
APPENDIX B Molybdenum - 8
References
AQUIRE (AOUatic Information REtrieval Database), 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth MN.
Arrington, L. R., and G. K. Davis. 1953. Molybdenum toxicity in the rabbit. Journal of
Nutrition 51:295-304.
Arrington, L. R., C. B. Ammerman, and J. E. Moore. 1965. Molybdenum toxicity in rats
and rabbits. Quart. J. of the Flor. Acad. of Sci. 28:129-136.
Arthur, D. 1965. Interrelationships of molybdenum and copper in the diet of the guinea pig.
J. Nutr. 51:295-304. As cited in U.S. EPA (Environmental Protection Agency). 1993c.
Integrated Risk Information System. April.
Cymbaluk, N.F., H.F. Schryver, H.F. Hintz, D.F. Smith and J.E. Lowe. 1981. Influence of
dietary molybdenum on copper metabolism in ponies. JMutr. 111:96-106, 1981.
i
Eisler,R. 1989. Molybdenum hazards to fish.wildlife, and invertebrates:a synoptic review.
U.S. Fish Wild. Serv. Biol. Rep.%5 (1.19). 61pp.
57FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139). Draft
Report: A Cross-species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg 3/4/day.
Fungwe, T. V., F. Buddingh, D. S. Demick, C. D. Lox, M. T. Yang, and S. P. Yang. 1990.
The role of dietary molybdenum on estrous activity, fertility, reproduction and
molybdenum and copper enzyme activities of female rats. Nutrition Research 10:515-
524.
Gilani, S.H. and Y.Alibhai: 1990. Teratogenicity of metals to chick embryos. J. of
Toxicology and Environmental Health, 30:23-31.
Jeter, M. A., and G. K. Davis. 1954. The effect of dietary molybdenum upon growth,
hemoglobin, reproduction and lactation of rats. Journal of Nutrition 54:215-220. As
cited in U.S. EPA (Environmental Protection Agency). 1993c. Integrated Risk
Information System. April.
Kienholz, E.W. Effects of Environmental Molybdenum Levels Upon Wildlife, Molybdedum on
the Environment, V2.
August 1995
-------
APPENDIX B Molybdenum - 9
Luckey, T.D. and B. Venugopal. Metal toxicity in mammals (1): Physiologic and chemical
basis for metal toxicity. Plenum Press, N.Y.
McConnell, R. P. 2977. Toxicityv of molybdenum to rainbow trout under laboratory
conditions. Pages 725-730 in W. R. Chappell and K. K. Peterson (eds.). Molybdenum in
the environment. Vol. 2. The geochemistry, cycling, and industrial uses of molybdenum.
Marcel Dekker, New York.
Miller, R. F., N. O. Price, and R. W. Engel. 1956. Added dietary inorganic sulfate and its
effects on rats fed molybdenum. /. Nutr. 60:539-547. As cited in U.S. EPA
(Environmental Protection Agency). 1993c. Integrated Risk Information System. April.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. lexicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Pitt, M., J.Fraser and D.C. Thurley. 1980. Molybdenum toxicity in sheep, epiphysiolysis,
exotoses and biochemical changes. J.Compfath. 90:567T576.
Reid, B.L., A.A. Kurnick, R.L. Svacha and J.R.Couch. 1956. The effect of molybdenum on
chick and poult growth. Proc. Soc. Exp. Biol. Med. 93.
Ridgway, L.P. and D.A Kamofsky. 1952. The effects of metals on the chick embryo:
Toxicity and production of abnormalities in development. Ann. N.Y. Acad. Sci. 55:203.
Schroeder, H. A., and M. Mitchener. 1971. Toxic effects of trace elements on the
reproduction of mice and rats. Arch Environ Health 23:102-106.
i
Suter Di, G. W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects of Aquatic Biota: 1994 Revision. DE-AC05-
84QR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. EPA P338-179874. U.S.
EPA, Cincinnati, OH:
U.S. EPA (Environmental Protection Agency). 1990. Molybdenum: Drinking Water Health
Advisory Draft. Office of Water. September.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC. Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2):
Chemical toxicity of metals and metalloids. Plenum Press, N.Y., 1978.
August 1995
-------
APPENDIX B Molybdenum - 10
U.S. EPA (Environmental Protection Agency). 1993. Integrated Risk Information System.
April.
Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2): Chemical toxicity of metals
and metalloids. Plenum Press, N.Y., 1978.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
White,R.D., R.A. Swick, P.R. Cheeke. 1984. Effects of dietary copper and molybdenum on
tansy ragwort (Senecio jacobaea) toxicity in sheep. Am, J.Vet.Res. 45(1).
Wide, M. Effect of short-term exposure to five industrial metals on the embryonic and fetal
development of the mouse. 1984. Environmental Research 33:47-53.
Wittenberg, K.M. and T.J. Devlin. 1987. Effects of dietary molybdenum on productivity and
metabolic parameters of lactating beef cows and their offspring. Can. J. Anim. Sci.
67:1055-1066.
Wren, C.D., H.R. Maccrimmon and B.R. Loescher. 1983. Examination of bioaccumulation
and biomagnification of metals in a precambrian shield lake. Water, Air, Soil Pollution
19:277-291.
August 1995
-------
Freshwater Tox. < - Molybdenum
Cas No. 7439-98-7
Chemical
Name
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
Species
aquatic
organisms
daphnid
daphnid
rainbow trout
rainbow trout
rainbow trout
Type of
Effect
chronic
chronic
chronic
chronic
acute
acute
Description
scy
CV
EC20
NOEC
LC50
LC50
Value
239
880
360
17
1320
800
Units
ug/L
ug/L
ug/L
ppm
ppm
ppm
Test Type
(Static/Flow
Through)
NS
NS .
NS
NS
static
static
Exposure
Duration
/Timing
NS
NS
NS
1 year
96 hours
96 hours
Reference
Suter and Mabrey, 1 994
Suter and Mabrey, 1 994
Suter and Mabrey, 1994
McConnell, 1977
McConnell, 1977
McConnell, 1977
Comments
No significant biological
differences in mortality, growth or
hematocrits. Molybdenum did not
exert a toxic effect on eyed eggs,
sac-fry or fingerling stages of
development.
Trout averaging 55 mm in length.
Trout averaging 20 mm in length.
-------
Terrestrial Toxicity - Molybdenum
Cas No. 7439-98-7 .
Chemical
Name
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
molybdenum
Species
rat
guinea pig
rat
mouse
rat
rat
rat
rat
rabbit
rabbit
rabbit
rabbit
Type ot
Effect
growth
growth
NS
repro
repro
repro
growth
growth
growth
growth
growth
growth
Descripti
on
LOAEL
LOAEL
AEL
AEL
LOAEL
AEL
NOAEL
LOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
20
40
7.5
1.18
5.6
700
500
1000
1000
2000
0.05
0.1
Units
ppm
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-d
PPH!
ppm
PPm_.
PPm
PPm
%diet
% diet
Route
(oral,.
s.c., i.v.,
i.p..
injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
1 3 weeks
8 weeks
6 weeks
6 months
1 1 weeks
10 days
six weeks
six weeks
three weeks
three weeks
12 weeks
1 2 weeks
Reference
Jeter and Davis, 1954
Arthur, 1965 as cited in
IRIS, 1993
Miller et al.. 1956 as cited
in IRIS, 1993
Schroeder and Milchener,
J971
Jeter and Davis; 1954_
Jeter and Davis, 1954
Arlington et al., 1965
Arrington et al., 1965
Arringtpne^al., 1965
Arrington et al., 1965
Arrington and Davis, 1 953
Arrington and Davis, 1953
Comments
Retarded weight gain and depigmehtation
when fed molybdenum at this dose with a
supplement of 5 ppm copper..
Doses ranged from the basal diet to 320 mg
Mo/kg-day; beginning with a decrease in '
weight gain at the lowest dose, toxic effects
increased in severity as the dose was
increased.
The authors consider this dose level a
LOAEL, however, only two doses were given
and there was no specification of increasing
severity of effects at the higher dose (30
mg/kg-day).
Excess fetal mortality and infertility in the F3
generations.
males recieved Mo from weaning, effects-
decreased tt of liner as a result of male
sterility. Orig. dose 80 and 140 ppm - to
calc. per day used Body Wt = 5 kg and Food
intake=.035 kg/d (U.S.EPA, 1988)
Irregular esjrus cycles in female rats.
_,
Reduced voluntary feed intake, growth and
efficiency of feed utilization.
Reduced voluntary feed intake and growth.
Anorexia, loss qf weight,
alopecia.dermatosis, anemia and death.
Some rabbits exhibited deformities of the
front legs.
-------
Freshwater Biological Uptake Measures - Molybdenum
Cas No. 7439-98-7
Chemical
Name
Species
B-factor
(BCF, BAF,
BMP)
Value
Measured
or
Predicted
(m,p)
Units
Reference
Comments
-------
Terrestrial Tox. / - Molybdenum
Cas No. 7439-98-7
Chemical
Name
molybdenum
molybdenum
molybdenum
Species
rat
rat
mouse
Type of
Effect
fel
let
fet
Descripti
on
NOAEL
LOAEL
AEL
Value
1.03E-01
0.19
100
Units
mg/kg-day
mg/kg-day
mM
Route
(oral,
s.c., i.v.,
i-p..
injection)
oral
oral
f.v.
Exposure
Duration
/Timing
6 weeks prior
to mating thru
day 21 of
gestation
6 weeks prior
to mating thru
day 21 of
gestation
day 8 of
gestation
Reference
Fungweetal , 1990
Fungweelal , 1990
Wide, 1984
Comments
See profile for conversion calculations from
mg/L and mg/kg-diet.
Increases in the number of dams with
resorbed fetuses, increased sites of
resorption and decrease in the average litter
size. See profile for conversion calculations
from mg/L and mg/kg-diet.
Decreases in normal fetal weight gain and
skeletal ossification.
i
-------
TerrestrialBiological Upk. . Measures - Molybdenum
Cos No. 7439-98-7
Chemical
Name
silver
Species
whole-plant
B-factor
(BCF. BAF,
BMP)
BCF
Value
4.0 E-01
Measured
or
Predicted
(m^P).. .
m
units
(ug/g DW plantj/(ug/g
soil)
Reference
Baes et al, 1983
Comments
Molybdenum - Page
-------
APPENDIX B Nickel - 1
Toxicological Profile for Selected Ecological Receptors
Nickel
Cas No.: 7440-02-0
Summary: This profile on nickel summarizes the toxicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates, and invertebrates (e.g.,
earthworms). The entire toxicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from data
presented in the technical support document for the Hazardous Waste Indentification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwate
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C_) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several studies were identified which investigated the effects of nickel exposure on
mammalian species. In a 3-generation study, Schroeder et al. (1971, as cited in IRIS, 1994)
exposed mice and rats to 5 mg of nickel per liter of drinking water, corresponding to a daily dose
of 0.43 mg/kg-day. In all generations, there was an increase in young deaths and runts as well
as decreases in litter size. There was also an increase in the proportion of males born. Smith
et al. (1990, as cited in IRIS, 1994) exposed rats to nickel in doses of 1.3, 6.8, and 31.6 mg/kg-
day for an 11-week pre-mating period. In the first generation, the proportion of dead pups per
litter increased for those groups given 31.6 mg/kg-day. However, the same elevation in dead
pups per litter was also seen in the second generation for those groups given 1.3 mg/kg-day and
6.8 mg/kg-day. In a 3-generation study conducted by Ambrose et al. (1976), rats were exposed
to 250, 500 and 1000 ppm of dietary nickel. The average weaning body weight was adversly
effected in weanlings of females on the 1000 ppm diet. This resulted in a LOAEL of 1000 ppm
August 1995
-------
APPENDIX B Nickel-2
and a NOAEL of 500 ppm for this developmental effect. The NOAEL from the Ambrose et al.
(1976) study was converted to a daily dose of 54.13 mg/kg-day for the purpose of calculating
benchmarks. The bodyweight and the food intake rate of the test species were needed for this
conversion. The bodyweight of the male and female rats was estimated by using the geometric
mean of the weights presented in the study's control groups at weeks 1, 3, 6, and 13. The food
intake rate was determined by using the food consumption equation for laboratory mammals
(Nagy, 1987).
The NOAEL for developmental effects from the Ambrose et al. (1976) study was chosen to
derive the lexicological benchmark because (1) chronic exposures were administered via oral
ingestion, (2) it focused on irregularities in the development of offspring as a critical endpoint,
(3) the study contained dose response information, and (4) the study reported the lowest toxicity
value for a critical endpoint The study by Schroeder et al. (1973) was not selected for the
derivation of a benchmark due to the administration of only one test dose resulting in a lack of
appropriate dose-response information. The Smith et al. (1990) study was not chosen due to
confounding dose-response information presented in the study.
The study value from Ambrose et al. (1976) was scaled for species representative of a freshwater
ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994):
Benchmark = NOAEL. x
where NOAELt is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 PR 24152). Since the Ambrose et al.
(1976) study documented developmental effects resulting from nickel exposure to male and
female rats, the mean of the male and female body weights of each representative species was
used in the scaling algorithm to obtain the toxicological benchmarks.
Data were available on the reproductive and developmental effects of nickel, as well as growth
or chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. The data set contained a study
value for reproductive endpoints (Schroeder et al., 1971, as cited in IRIS, 1994) that was more
than an order of magnitude below the benchmark value. Based on the data, set for nickel the
benchmarks developed from the Ambrose et al. (1976) study were categorized adequate, with
a "*" to indicate that some adverse effects have been observed at the benchmark level.
Birds: No suitable studies were found for nickel toxicity in avian species associated with the
freshwater ecosystem.
August 1995
-------
APPENDIX B Nickel - 3
Fish and aquatic invertebrates: The Final Chronic Value (FCV) for nickel of 1.6E-01 mg/L was
selected as the benchmark protective of fish and aquatic invertebrates (U.S. EPA, 1985). The
FCV for nickel is based on the equation e*0-846^1"^1*")!-1-1645). It is a hardness dependent
criterion that is normalized to 100 mg/L. Since the FCV was derived in the AWQC document,
the benchmark was categorized as adequate, with a "*" to indicate that adverse effects have been
seen at or below the benchmark value.
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEQ for vascular aquatic
plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastrum capricornutum). For nickel the benchmark
value presented in Suter and Mabrey (1994) of 5.0E-03 mg/L was based on the incipient
inhibition of Micrdcystis aeruginosa. As described in Section 4.3.6, all benchmarks for aquatic'
plants were designated as interim.
Benthic community: The nickel benchmark protective of benthic organisms is pending a U.S. EPA
review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Nickel - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
fijMMiWI
mink .
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring guN
kingfisher
BenoNwirtc
¥«kM"«NQft*
to*
33.41
19.97
ID
ID
ID
ID
ID
ID
ID
ID
ttu*
SlMKlMt
rat
rat
-
-
-
-
-
ifteel
dev
dev
-
-
-
-.
•
*&*+*
54.13
54.13
-
-
•
-
-
*-*.
NOAEL
NOAEL
. -
•
-
• - .
Of
+
•
•
-
• • ,
•
-
-
%
OriQjMfSowt*
Ambrose at al., 1976
Ambrose eta!., 1976
-
.-
•
-
-
-
-
-
'Benchmark Category, a - adequate, p = provisional, i = interim; a '*' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = insufficient data
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
W^w*w^^JWIWI^^
ftpecihNi
fish and aquatic
invertebrates
aquatic plants
benthic community
wW^fcrt^nt
Vato**
«flfc
1.6E-01 (a')
5.0E-00
under review
*****
aquatic
organisms
aquatic •
plants
~*.
FCV
CV
OrityratftmtoD*
AWQC (hardness
dependent)
Suter and Mabrey,
1994
•
'Benchmark Category, a = adequate, p = provisional, i = interim; a "" indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Nickel - 5
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to nickel.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Ambrose et al., 1976) was used to derive the nickel lexicological benchmark for
mammalian species representing the terrestrial ecosystem. The study NOAEL of 54.13
mg/kg-day was scaled for representative species in the terrestrial ecosystem using a cross-
species scaling algorithm adapted from Opresko et al. (1994). Since the .Ambrose et al.
(1976) study documented developmental effects resulting from nickel exposure to male and
female rats, the mean of the male and female body weights of each representative species was
used in the scaling algorithm to obtain the lexicological benchmarks.
Based on the data set for lead the benchmarks developed from die Ambrose et al. (1976)
study were categorized as adequate, with a "*" to indicate that some adverse effects have
been observed at the benchmark level.
Birds: No suitable studies were found for nickel toxicity in avian species associated with the
terrestrial ecosystem.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the LOEC values and then approximating the 10 percentile.
If there were 10 or fewer values, the 10th percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation. The
selected benchmark for phytotoxic effects of nickel in soils is 30 mg/kg (Will and Suter,
1994). Since the study value selected is the 10th percentile of more than 10 LOEC values, the
terrestrial benchmark for lead is categorized as provisional.
Soil Community: Adequate data with which lo derive a benchmark protective of ihe soil
community were nol identified.
August 1995
-------
APPENDIX B
Nickel - 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
nsymntatfr*
*»«(•«
daar mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
toil community
Sencbtnarii
Vafti**
*9***y
90.43
92.98
78.77
31.96
23.01
21.81
11.06
ID
ID
10
ID
ID
30(p)mg/kg
ID
Study
dfMde*
rat
rat
rat
rat
rat
rat
rat
>"
-
-
terrestrial
plants
• •
iftaet
dev
dev
dev
dev
dev
dev
dev
-
-
-
growth/
yield
-
SHM^
ttfe*
«***>
**t
54.13
54.13
54.13
54.13
54.13
54.13
54.13
-
-
-
-
-
30 mg/Vg
-
H^Mffolfefe
^^^^^F*J^P^W»P
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
•
-
-
•
-
10th percantile
LOEC
.
3F
-
-
-
-
-
•
-
-
-
•
-
- •
< \
OdelMlSw.ro* :
Ambrose et al., 1976
Ambrose et al.. 1976
Ambrose et at., 1976
Ambrose etal., 1976
Ambrose etal., 1976
Ambrose et al., 1976
Ambrose et al., 1976
'
• '
-
-
-
Will and Suter, 1994
-
•Benchmark Category, a =• adequate, p = provisional, i = interim; a "" indicate* that the benchmark value was an order
of magnitude or more above tw NEL or LEI for other adverse effects.
ID » insufficient data
August 1995
-------
APPENDIX B Nickel - 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. For
metals, BCFs are whole-body bioconcentration factors and refer to total surface water
concentrations (versus freely dissolved concentrations). Consequently all calculations of
acceptable tissue concentrations (TC) represent whole-body concentrations. The following
discussion describes the rationale for selecting the biological uptake factors and provides the
context for interpreting the biological uptake values.
The single bioconcentration factor for muscle suggested that nickel does not bioconcentrate in
fish. However, lacking data on whole-body bioconcentration, this value should be intepreted
with caution. The BCF value for fish was a measured value of 1.00 (Stephan, 1993).
Insufficient data were identified on bioconcentration in aquatic invertebrates. Appropriate
studies on bioaccumulation/bioconcentration were not identified for terrestrial vertebrates and
invertebrates (including earthworms). A whole plant-BCF value of 1.1E-01 was presented by
U.S.EPA (1992e). For metals, empirical data were used to derive the BCF for aboveground
forage grasses and leafy vegetables. In particular, the uptake-response slope for forage
grasses was used as the BCF for plants in the terrestrial ecosystem since most of the
representative plant-eating species feed on wild grasses.
August 1995
-------
APPENDIX B
Nickel - 8
Table 4. Biological Uptake Properties
. «GOfegfa»<
fish
(Moral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BAF.or
BSAF
BCF
*
•
•
.
BCF
tiflkHMWKlM'
whoU>bo
-------
Terrestrial Toxicity - Nickel
Cas No. 7440-02-0
Chemical
Name •
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
Species
rat
rat
rat
rat
dog
dog
rat
rat
rat
Type of
Effect
growth
growth
dev
dev
growth
growth
growth
growth
tet
Description
NOAEL
LOAEL
LOAEL
NOAEL
NOAEL
LOAEL
NOAEL
LOAEL
LOAEL
Value
5
50
108.26
54.1
25
63
5
35
51.6
. Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day;
mg/kg-day
mg/kg-day
Exposure
Route (oral,
s.c.. i.v., i.p.,
injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure Duration
timing
2 years
2 years
3-generation
3-generation
NS H
NS
90 days
90 days
90 days prior to mating
Reference
Ambrose et al., 1976
Ambrose^al. , 1 976
Ambrose et al., 1976
Ambrose et al., 1976
Ambrose et al., 1976
Ambrose et al., 1976
ABC, 1986 as cited in IRIS,
1994
ABC, 1986 as cited in IRIS,
1994
RTI, 1987 as cited in IRIS,
1994
Comments
Dose (mg/kg-d) was estimated
by|RIS
Depressed body weight gain.
Dose (mg/kg-d) was estimated
by IRIS
Doses were 0.250,500,1000
ppm. Decrease in average
weaning body wt.. Dose
converted from 1000 ppm using
26 week body wt. = 309 g, and
food intake=0.03 kg/d (USEPA,
1988)
Dose converted from 500 ppm
using 26 week body wt. = 309
g, and food intake=0.03 kg/d
(USEPA, 1988)
Doses were 0,100,1000,2500
ppm. Dose (mg/kg-d) was
estimated by IRIS and 100
ppm.
Depressed body weight gain.
Dose (mg/kg-d) was estimated
by IRIS
Doses were 0,5,35, and 100
mg/kg-d
Decreased body and organ
weights; slight signs of toxicity
such as lethargy, ataxia,
irregular breathing and
discolored extremities.
F1 & F2 generation- increased
pup mortality and decreased
pup weight. Effects are ?
beacuse food/water intake was
sign, lower in the two higher
dose groups as compared to
controls.
-------
APPENDIX B Nickel - 11
Suter n, G.W. and J.B. Mabrey. 1994. Toxirological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
Syracuse Research Corporation. 1986. Toxicological Profile for Nickel. Prepared for
Agency for Toxic Substances and Disease Registry (ATSDR), U.S. Public Health Service,
in collaberation with U.S. Environmental Protection Agency.
U.S. EPA (Environmental Protection Agency). 1985. Health Effects Assessment for Nickel.
U.S. Environmental Protection Agency Rep. 600/8-83/012F. Available from NTIS, 5285
Port Royal Road, Springfield, VA 22161.
U.S. EPA (Environmental Protection Agency). 19881. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. EPA P338-179874. U.S.
EPA, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1992. TSC1292 Criteria Chart. 304(a)
Criteria and Related Information for Toxic Pollutants. Water Management Division,
Region IV.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993. Integrated Risk Information System.
June.
Watras, C. J., J. MacFarlane, and F. M. M. Morel. 1985. Nickel accumulation by
Scenedesmus and Daphnia: Food-chain transport and geochemical implications. Can. J.
Fish. Aquat. Sci. 42:724-730.
Will, M. E., and G. W. Suter, II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial . ..city - Nickel
Cas No. 7440-02-0
Chemical
Name
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
Species
rat
rat
rat
rat
rat
rat
rat
rat
mouse
mouse
mouse
rat
Type of
Effect
fet
liver
liver
fet
rep
rep
emb
emb
rep
rep
emb
fet
Description
AEL
NOAEL____
LOAEL
AEL
LOAEL
NOAEL
LOAEL
NOAEL
AEL
AEL
AEL
AEL
Value
7.3
30.8
51.6
31.6
50
50
12
8
20
20
20
0.598
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg-d
Exposure
Route (oral,
s.c., i.v., i.p.,
injection)
oral
oral
oral
oral
oral
oral
i.m.
i.m.
' P
'P
'P
oral
Exposure Duration
/Timing .
90 days prior to mating
90 days prior to mating
90 days prior to mating
1 1 weeks prior to
mating
Fo-11 weeks; F1and
F2- 105 days
Fo-11 weeks; Fland
F2- 105 days
Day 8 of gestation
Day 8 of gestation
Day 1 of gestation
Days 1 -6 of gestation
Day 1 of gestation
3 generations
Reference
RTI, 1987 as cited in IRIS,
1994
RTI, 1987 as cited in IRIS,
1994
RTI, 1987 as cited in IRIS,
1994
Smith et al., 1990 as cited
in IRIS. 1994
Ambrose el al, 1976 and
Borzelleca et al., 1965 as
cited in ATSDR, 1993
Ambrose etal, 1976 and
Borzelleca et al., 1965 as
cited in ATSDR, 1993
Sunderman et al., 1978
Sunderman et al., 1978
Storeng and Jonsen, 1981
Storeng and Jonsen, 1981
Storeng and Jonsen; 1981
Schroeder el al., 1971
Comments
These effects were not
considered to be significant by
the IRIS editors since they were
not seen in both the higher
dose groups. ,
doses were 0,50,250,500 ppm.
NOAEL=250 ppm
Decrease in maternal body
weight; decreases in absolute
and relative liver weights.
Elevated number of dead pups
per litter. Several doses (0, 1 .3,
6.8 and 31.6) were used-effects
do not exhibit a clear dose
response relationship.
Decreased offspring per litter.
No effect on fertility, gestational
length or viability.
Increased ratio of dead/live
fetuses.
Significant decrease in the
number of implantations.
Increased frequency of
resorptions.
Increased number of abnormal
and stillborn fetuses. '
Decreased litter size, increases
in young deaths and runts.
-------
Freshwater Toxicity - Nickel
Cas No. 7440-02-0
Chemical
Name
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
Species
fathead
minnow
fathead
minnow
aquatic
organisms
fish
daphnid
fish
daphnid
rock bass
Daphnia
pulicaria
daphnid
daphnid
daphnid
fathead
minnow
Type of
Effect.
rep
rep
chronic
chronic
chronic
chronic
chronic
acute
acute
acute
acute
chronic
acute
Description
NOEC
LOEC
NAWQC
LOEC
LOEC
EC20
EC20
LC50
LC50
LC50
LC50
LOEC
LC50
Value
380
730
160
35
5
62
45
697-3757
(2154)
2182
510
1120
30
2916-17678
(6248)
Units
ug/l ^
ug/l
ug/L
ug/L
ug/L
ug/L
y»A
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS__
NS
Exposure
Duration
/Timing
life-cycle
life-cycle
NS
NS
NS
NS
NS
96 hours
48 hours
48 hours
48 hours
3 weeks
96 hours
Reference
Dickering, 1974
Pickering, 1974
40C.F.R. Part 131, Vol.
57, No. 246, p. 60848
(1992)
Nebeker, 1985
Lazareva, 1985 as cited
in Suteretal., 1992
Suteretal., 1992
Suteretal., 1992
Lindetal., 1978 as cited
inAQUIRE, 1994
Lindetal., 1978 as cited
inAQUIRE. 1994
Biesinger & Christensen,
1972
Biesinger & Christensen,
1972
Biesinger & Christensen,
1972
Lindetal., 1978 as cited
in AQUIRE, 1994
Comments
No adverse effects on survival,
growth and reproduction
Statistically significant effect on
the number of eggs produced per
spawning and the hatchability of
these eggs.
Without food.
With food.
/•
16% reproductive impairment.
526.687763
i
-------
APPENDIX B Nickel - 9
References
ABC (American Biogenics Corp.). 1986. Ninety-day gavage study in albino rats using
nickel. Draft Final Report submitted to Research Triangle Institute, P.O. Box 12194,
Research Triangle Park, NC 27709. As cited in U.S. EPA (Environmental Protection
Agency). IRIS (Integrated Risk Information System). 1994.
Ambrose, A. M., P. S. Larson, J. F. Borzelleca, and G. R. Hennigar, Jr. 1976. Long term
toxicologic assessment of nickel in rats and dogs. /. Food Sc. & Tech. 13:181-187.
AQUIRE (AQUatic Toxicity /nformation flEtrieval Database). 1994. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Biesinger, K. E., and G. M: Christensen. 1972. Effects of metals on survival, growth,
reproduction, and metabolism of Daphnia magna. J. Fisheries Res. Bd. of Canada
28:1691-1700.
Calamari, D., G. F. Gaggino, and G. Pacchetti. 1982. Toxicokinetics of low levels of Cd,
Cr, Ni and their mixture in long-term treatment on Salmo gairdneri rich. Chemosphere
57 FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg3/4/day.
Lazareva, L.P. 1985. Changes in biological characteristics of Daphnia magna from chronic
action of copper and nickel at low concentrations. Hydrobiol. J. 21(5): 59-62.
Lind, D., K. Alto, and S. Chatterton. 1978. Regional Copper-Nickel Study Draft Report.
Minnesota Environmental Quality Board, St. Paul, MN. 54 pp. As cited in AQUIRE
(AQUatic Toxicity Information REtrieval Database). Environmental Research Laboratory,
Office of Research and Development, U.S. Environmental Protection Agency, Duluth,
MN.
Mathur, A. K., K. K. Datta, S. K. Tandon, T. S. S. Dikshith. 1977. Effect of nickel sulfate
on male rats. Bulletin of Environmental Contamination and Toxicology. 17:241-248.
Nagy, K. A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol Mono 57:111-128.
Nebeker, A.V., C. Savonen, and D.G. Stevens. 1985. Sensitivity of rainbow trout early life
stages to nickel chloride. Environmental Toxicology and Chemistry. 4: 233-239.
August 1995
-------
APPENDIX B Nickel - 10
Opresko D. M., B. E. Sample, and G.W. Suter II. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/Rl.
Pickering, Q. H. 1974. Chronic toxicity of nickel to the fathead minnow. Journal WPCF
46:760-765.
RTI (Research Triangle Institute). 1987. Two generation reproduction and fertility study of
nickel chloride administered to Cd rats in drinking water: Fertility and reproductive
performance of the Po generation (Pan II of HI) and Fl generation (Part ni of HI). Final
study report Report submitted to Office of Solid Waste Management, U.S. EPA,
Washington, DC. As cited in As cited in U.S. EPA (Environmental Protection Agency).
IRIS (/ntegrated Risk /nformation System). 1994.
Schroeder, H. A., and M. Mitchener. 1971. Toxic effects of trace elements on the
reproduction of mice and rats. Arch Environ Health. 23:102-106.
Smith, M. K., J. A. George, H. F. Stober, and G. L. Kimmel. 1990. Perinatal toxicity
associated with nickel chloride exposure. Fund. Appl. Toxicol. Preliminary unpublished
draft. As cited in U.S. EPA (Environmental Protection Agency). IRIS (/ntegrated Risk
/nformation System). 1994.
Sreedevi, P., A. Suresh, B. Sivaramakrishna, B. Prabhavathi, and K. Radhakrishnaiah. 1992.
Bioaccumulation of nickel in the organs of the freshwater fish, Cyprinus carpio, and the
freshwater mussel, Lamellidens marginalis, under lethal and sublethal nickel stress.
Chemosphere. 24:29-36.
Stephan, C. E. 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and DEvelopment, Duluth, MN, PB93-154672.
Storeng, R., and J. Jonsen. 1981. Nickel toxicity in early embryogenesis in mice.
Toxicology 20:45-51.
Sunderman, Jr., F. W., S..K. Shen, J. M. Mitchell, P. R. Allpass, and I. Damjanov. 1978.
Embryotoxicity and fetal toxicity of nickel in rats. Toxicology and Applied Pharmacology
43:381-390.
Sunderman, F.W., S. K. Shen, M. C. Reid, P.R. Allpass. 1980. Teratogenicity and
embryotoxicity of nickel carbonyl in Syrian hamsters. Teratogenesis, Carcinogenesis, and
Mutagenesis. 1:223-233.
August 1995
-------
Freshwater Biological ,_iake Measures - Nickel
Cos No. 7440-02-2
Chemical
Name
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
nickel
Species
Salmo
gairdneri
Salmo
gairdneri
Salmo.
gairdneri
Salmo '
gairdneri
Salmo
gairdneri
Salmo
gairdneri
fish
rainbow trout
daphnid
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
4.2
4.5
3.1
3.2
0.9
1.1
47
0.8
2-12
Measured
. or
Predicted
(m,p)
P
P
P
P
P .... .
P
m
m
m
Units
NS
NS
NS
NS
NS
NS
L/Kg
NS
NS
Reference
Calamari et al., 1982
Calamari et al., 1982
Calamari et al., 1982
Calamari et al.. 1982
Calamari et al., 1982
Calamari etal., 1982
U. S. EPA, 1992
Calamari et al., 1982 as
cited in U. S. EPA. 1993b
Watras etal., 1985
Comments
Predicted BCF for the kidney based
exposure to 1 mg Ni/L.
Predicted BCF for the kidney based
exposure to 1mg Ni/L.
Predicted BCF for the liver based
exposure to 1 mg Ni/L.
Predicted BCF for the liver based
exposure to 1mg Ni/L.
Predicted BCF for the muscle based
exposure to 1mg Ni/L.
Predicted BCF for the muscle based
exposure to Img Ni/L.
Normalized to 3% lipid.
Muscle BCF.
-------
Terrestrial Biological Uptake Measures - Nickel
Cas No. 7440-02-0
Chemical
Name .
nickel
Species
plant
B-lactor
(BCF. BAF,
BMP)
BCF
Value
0.022 -
Measured
or.
Predicted
(m,p)
P
units
(ug/g DW plant)/(ug/g
soil)
Reference
U.S. EPA, 1990 as cited in RTI,
1994
Comments
-------
APPENDIX B Parathion - 1
lexicological Profile for Selected Ecological Receptors
Parathion
Cas No.: 56-38-2
Summary: This profile on parathion summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater, ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include, terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
data presented in the technical support document for the Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ro) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with ihe freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife which
reported dose-response data.'However, lexicological studies involving parathion exposure to
mammals have been conducted using laboratory rats and mice. Fetoloxicily was observed in
pregnanl mice (Harbison, 1975) administered paraihion (dissolved in corn oil) via intraperitoneal
injection ai concentrations of 4, 8, 10, 11, and 12 mg/kg-day on days 8 to 14 of gestation. Based
on the number of prenatal deaths, Hardison recorded a LOAEL of 12 mg/kg-day. A chronic
reproductive study was identified in which weanling albino rats were fed dieiary concentrations
of 10, 20, 50, 75, or 100 ppm of parathion before mating through to panurition (Barnes and
Denz, 1951). Based on ihe two indices of fertility and viability of young, a NOAEL of 10 ppm
and a LOAEL of 20 ppm were chosen from the Barnes and Denz (1951) sludy. Based on ihe
reference body weighi (kg) and allometric equation for food consumption (kg/day) for rats
August 1995
-------
APPENDIX B Parathion - 2
reported in Recommendation for and Documentation of Biological Values for use in Risk
Assessment (U.S. EPA, 1988), the NOAEL was converted to 0.76 mg/kg-day and the LOAEL to
1.52 mg/kg-day. Deskin et al. (1979) investigated cholinesterase activity in perinatal rats
administered parathion doses of 0.01, 0.10, and 1 mg/kg-day via oral gavage from day 2 of
gestation to day 15 of the lactation period. Deskin et al., (1979) noted that female perinatal rats
showed a significant reduction in pseudo cholinesterase at doses of parathion at 0.01 mg/kg-day
(LOAEL).
The NOAEL in the Barnes and Denz (1951) study was selected as the study to derive the
lexicological benchmark because reproductive toxicity was the critical endpoint, dietary
concentrations were administered via oral ingestion during a critical life-stage period, and there
was sufficient dose-response information. The study by Deskin et al., (1979) was not selected
because the changes in cholinesterase activity observed were not indicative of population
effecting endpoints. The study by Harbison (1975) was not selected because of the uncertainity
associated with extrapolating injection exposure to wildlife exposure.
The NOAEL of 0.76 mg/kg-d (10 ppm) reported by Barnes and Denz (1951), was scaled for
species representative of a freshwater ecosystem using a cross-species scaling algorithm adapted
from Opresko et al. (1994)
( bw V4
Benchmark = NOAEL. x '-
•• IKJ
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57FR 24152). Since the benchmark study
documented reproductive effects from parathion exposure to male and female rats, the mean of
.female and male body weights for each representative wildlife species were used in the scaling
algorithm.
Data were available on the reproductive and developmental, effects of parathion, as well as
chronic survival. In addition, the data set contained studies which were conducted over chronic
and subchronic durations and during sensitive life stages. The identified studies were not
conducted using a range of wildlife species and as such, inter-species toxicity differences were
not identifiable. Therefore, an inter-species uncertainty factor was not applied. There was one
other mammalian value in the data set which reported a LOAEL for an endocrine endpoint that
was more than a magnitude lower than the benchmark value. Based on the data set for parathion,
the benchmarks developed from Barnes and Denz (1951) were categorized as adequate, with a
"*" to indicate that adverse effects may occur at the benchmark level.
Birds: No suitable chronic studies were found for representative avian species in which dose-
response data were reported. However, subchronic toxicity studies involving parathion have been
conducted using bobwhite quail. A subchronic reproductive study was identified in which thirty-
August 1995
-------
APPENDIX B Parathion-3
week old female bobwhite quails (Rattner et al., 1982) were fed a dietary concentration of 25 or
100 ppm of parathion (suspended in com oil) for 12 days. Rattner et al. observed reproductive
endpoints and tolerance to cold through the following parameters: body temperature, egg
production, and the plasma concentrations of luteinizing hormone, progesterone, and
corticosterone. Rattner et al. (1982) reported a NOAEL of 25 ppm and a LOAEL of 100 ppm.
Based on the reference body weight estimated by Roseberry and Klimistra (1971) and an
allometric equation for estimating daily food ingestion (Nagy, 1987), the NOAEL of 25 ppm was
converted to a daily dose of 2.7 mg/kg-day.
Because the study by Rattner et al., (1982) focused on reproductive toxicity as a critical endpoint
and dietary concentrations were administered via oral ingestion during a critical life-stage period,
the NOAEL of 2.7 mg/kg-d was chosen to derive the toxicological benchmark for freshwater
birds. The principles for allometric scaling were assumed to apply to birds, although specific
studies supporting allometric scaling for avian species were not identified. Thus, for the avian
species represented in the generic freshwater ecosystem, the NOAEL of 2.73 mg/kg-day from the
Rattner et al. (1982) study was scaled using the cross-species scaling method of Opresko et al.
(1994). Since Rattner et al. (1982) documented reproductive effects from parathion on female .
bobwhite quail, female body weights for each representative species were used in the scaling
algorithm to obtain the toxicological benchmarks.
The data set for the effects of parathion on birds was not extensive. Data were not available on
developmental endpoints, as well as, growth or survival effects. Subchronic laboratory
experiments were not conducted on a range of avian species and as such, inter-species differences
among wildlife species were not identifiable. Based on the limited avian data set for parathion,
the benchmarks developed were categorized as interim.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 1,3 E-5 mg/1 developed for
the AWQC for parathion was the lowest value recorded in the dataset (51 FR 43667). As such,
the FCV was selected as.the benchmark for fish and aquatic invertebrates in the generic
freshwater ecosystem. Because the FCV was developed for an AWQC, this benchmark was
categorized as adequate.
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum). Aquatic
plant data was not identified for parathion and, therefore, no benchmark was developed..
Benthic community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbon partition coefficient (K^) to determine a protective sediment concentration (Stephan,
1993). The EQp number is the chemical concentration that may be present in sediment while still
protecting the benthic community from the harmful effects of chemical exposure. The FCV for
parathion was used to calculate a EQp number of 7.24E-02 mg parathion /kg organic carbon.
August 1995
-------
APPENDIX B Parathion - 4
Assuming a mass fraction of organic carbon for the sediment (foc) of 0.05, the benchmark for the
benthic community is 3.62E-03 mg parathion /kg of sediment. Since the EQp number was based
on a FCV, the sediment benchmark was categorized as adequate.
August 1995
-------
APPENDIX B
Parathion • 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
ft*pfwwn1*tiV«
Spe$l»«
mink
river otter
bald eagle
osprey
great blue heron
mallard
lessor scaup
spotted
sandpiper
herring gull
kingfisher
Benchmark
v«fu»* jn$*g*
^
0.6 (a')
0.36 (a*)
1.2(0
1.5(0
1.4(0
1.7(i)
1.8(0
3.8 (i)
1.8(1)
2.8 (i)
Study Specte
rat
rat
bobwhite quail
bobwhite quail
bpbwhite quail
bobwhite quail
bobwhite quail
bobwhite quail
bobwhite quail
bobwhite quail
E««ct
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
Study
Value
mg/kg«Sfty
0.76
0.76
2.7
2.7
2.7
2.7
2.7
2.7
2.7
2.7
o«*crfpaoo
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
-
•
-
-
. •
•
-
•
Orfgto*t$o«re*
Barnes and Oenz,
1951
Barnes and Denz,
1951
Rattner et al., 1982
Rattneretal.. 1982
Rattner et al.. 1982
X
Rattner et al., 1982
Rattner et al.. 1982
Rattneretal.. 1982
Rattneretal., 1982
Rattneretal., 1982
'Benchmark Category, a - adequate, p « provisional, i = interim; a "" indicates that the benchmark value was ah order of
magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Parathion - 6
Table'2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
R«pr»««ntativ
• e Specie*
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark
V«to»*
wg/i
1.3E-05(a)
ID
3.6E-03 (a)
mg/kg sediment
Study Sp«ef«*
AWQC
organisms
-
benlhic
community
Descrtpfioa
FCV
-
FCV x K,,.
Onfltnal
Sourc*
51 FR 43667
51 FR 43667
IL
'Benchmark Category, a * adequate, p = provisional, i = interim; a "" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEL for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rational behind Toxicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to parathion.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Barnes and Denz, 1951) was used to derive the parathion lexicological benchmark for
mammalian species representing the terrestrial ecosystem. The NOAEL of 0.76 mg/kg-d
reported by Barnes and Denz (1951), was scaled for each of the species in the generic
terrestrial ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994). Since the benchmark study documented reproductive effects from parathion exposure
to male and female rats, the mean of female and male body weights for each representative
species was used in the scaling algorithm. Because the study value selected was a NOAEL
that was at least an order of magnitude above the lowest LOAEL in the data set, the
benchmarks developed from the Barnes and Denz (1951) study were categorized as adequate,
with a "*" to indicate that adverse effects may occur at the benchmark level..
Birds: No additional avian toxicity studies were identified for species representing the
terrestrial ecosystem. The NOAEL selected for the representative species of the generic
terrestrial ecosystem was 2.73 mg/kg-day form the study by Rattner et al. (1992). The
NOAEL was then scaled for the representative species using the cross-species scaling
August 1995
-------
APPENDIX B Parathion - 7
algorithm adapted from Opresko et al. (1994). Since the Rattner et al. (1956) study
documented reproductive effects from aldrin on female bobwhite quails, female body weights
for each representative species were used in ihe scaling algorithm to obtain the lexicological
benchmarks. Based on the limited avian data set for parathion, the benchmarks developed
were categorized as interim.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for parathion and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were riot available. Only one study (van Gestel.et al., 1992) which addressed the
developmental or reproductive effects of parathion on soil invertebrates was located in the
literature.
August 1995
-------
APPENDIX B
Parathion - 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
fteprMwitaQy*
$f«&* t
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern bobwhite
American robin
American
woodcock
plants
soil
Benchmark
VaJu»*
Wj|/k0-
-------
APPENDIX B
Parathion - 9
in. Biological Uptake Measures
This section presents biological uptake measures (i.e, BCFs, BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for selected ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For the generic aquatic ecosystems^ the BCF value is
identified as whole-body or lipid-based and designated with a "d" if the value reflects
dissolved water concentrations, and a "t" if the value reflects total surface water
concentrations. For organic chemicals with log K<,w values below 4, bioconcentration factors
(BCFs) in fish were always assumed to refer to dissolved water concentrations (i.e., dissolved
water concentration equals total water concentration). The brief discussion following Table 4
describes the rationale for selecting the biological uptake factors and provides the context for
interpreting the biological uptake values.
Table 4. Biological Uptake Properties
«c«to0lc«J
iwcaptor
fish
trophic level 2
invertebrate*
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plant*
BCF.8AF.or
BSAF
BCF
BCF
BAF
BCF
BCF
BCF
tfpfM«iQr
whd*bo«ty
lipid
lipid
whole- body
whole-body
whole-body
whole-plant
v«tu»
6,325 (d or t)
•
7.9E-05,
7.6E-05
6.1E-04
0.24
»ourc* .
predicted value based on •
Thomann, 1080
insufficient data
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDO
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDO
U.S. EPA, 19920
d - refers to dissolved surface water concentration
t « refers to total surface water concentration
The bioconcentration factor for fish was estimated from the Thomann (1989) model (i.e., log
Kow - dissolved BCF/) because: (1) no appropriate measured values were identified, (2) the
BCF was in close agreement with predicted BCFs based on other methods (i.e., regression
equations), and (3) there were no data (e.g., metabolism) to suggest that the log K,,,
relationship deviates for parathion (log K,JW = 3.81). As stated in section 5.3.2, the dissolved
bioconcentration factor (BCF* ) for organic chemicals with log Kow below 4 was considered
BCF;d
August 1995
-------
APPENDIX B Parathion - 10
to be equivalent to the total bioconcentration factor (BCF/) and, therefore, adjusting the
BCF/1 by the dissolved fraction (fd) was not necessary.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial
wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial
Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient
quality to serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking
measured data is compared to the BBF for TCDD and that ratio (i.e., parathion BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B Parathion-11
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
AQUIRE C40C/atic Toxicity Mormation £Etrieval Database), 1995. Environmental
Research Laboratory, Office of Research and Development, U:S. Environmental Protection
Agency, Duluth,
Barnes, J.M., and F.A. Denz. 1951. The chronic toxicity of p-njtrophenyl diethyl
thiophosphate (E. 605). A long-term feeding experiment with rats. J. of Hygiene
49:430-441.
Deskin, R, L. Rosenstein, N. Rogers, and B. Westbrook. 1979. Parathion toxicity in
perinatal rats exposed in utero. Toxicol. Letters 3:11-14.
Dortland, R.J. 1980. Toxicological Evaluation of Parathion and Azinphosmethyl in
Freshwater Model Ecosystems. Versl. Landbouwkd. Onderz, 898:1-1.12. As cited in
AQUIRE (AOUatic Toxicity /nformation /?Etrieval Database), 1995. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Fleming, W.J., H. de Chacin, O.K. Pattee, and T.G. Lamont 1982. Parathion Accumulation
in Cricket Frogs and its Effect on American Kestrels. Journal of Toxicology and
Environmental Health, 10:921-927.
Harbison, R.D. 1975. Parathion-induced toxicity and ptienobarbital-induced protection
against parathion during prenatal development. Toxicol. Appl. Pharmacol. 32:482-493.
Hoffman, D.J. and P.H. Albers. 1984. Evaluation bf Potential Embryotoxicity and
Teratogenicity of 42 Herbicides, Insecticides, and Petroleum Contaminants to Mallard
Eggs. Archives of Environmental Contamination and Toxicology, 13:15-27.
Kimbrough, R.D., and T.B. Gaines. 1968. Effect of organic phosphorus compounds and
alkylating agents on the rat fetus. Arch. Environ. Health 16:805-808.
Nagy, K.A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol.Mono. 57:111-128.
August 1995
-------
APPENDIX B Parathion - 12
NIOSH. 1976. Criteria for a Recommended Standard... Occupational Exposure to
Parathion. U.S. Deaprtment of Health, Education, and Welfare, Public Health Service,
Center for Disease Control, National Institute for Occupational Safety and Health.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. lexicological Benchmarks for'Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy,'Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Rattner, B.A., L. Sileo, and C.G, Scanes. 1982. Hormonal Responses and tolerance to cold
and female quail following parathion ingestion. Pesticide Biochemistry Physiol.
18:132-138.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Sanders, H.O. and O.B. Cope. 1966. Toxicities of Several Pesticides to Naiads of Three
Species of Stoneflies. Limnol. Oceanogr., 13(1):112-117. As cited in AQUIRE (AOU&tic
Toxicity /nformation /?Etrieval Database), 1995. Environmental Research Laboratory,
Office of Research and Development, U.S. Environmental Protection Agency, Duluth,
MN.
Spacie, A., A.G.Vilkas, G.F. Doebbler, W.J. Kuc, and G.R. Iwan. 1981. Acute and chronic
parathion toxicity to fish and invertebrates. Contract No. 68-01-0155, Manuscript Office
of Research and Monitoring, U.S. Environmental Protection Agency, Washington, DC.
As cited in AQUIRE (AOUztic Toxicity_/nformation /?Etrieval Database), Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Stephan, C.E., D.I. Mount, D.J. Hansen, J.H. Gentile, G.A. Chapman, and W.A. Brungs.
(1985) Guidelines for Deriving Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and their Uses. U.S. Environmental Protection Agency,
Office of Research and Development, Environmental Research Laboratories. NTIS No.
PB85-227049.
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter, G.W. and J.W. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-96/R1.
U.S. Department of Energy, Oak Ridge National Laboratory, Orak Ridge, TN.
Roseberry and Klimistra. 1971. Annual weight cycles in male and female bobwhite quail.
Auk 88:116-123.
August 1995
-------
Terrestrial Toxicity - Parathion
Cas No. 56-38-2
Chemical
Name
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
Species
female
Sherman
strain rats
pregnant
mice
adult
pregnant rats
weanling
albino rats
weanling
albino rats
pregnant rats
bobwhite
quail
bobwhite
quail
Endpolnt
terat
body wt.,
mortality
endocrine
rep
rep
rep, dvp
rep
rep
Description
AEL
LOAEL
LOAEL
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
Value
3
4
0.01
0.76
1.52
1.5
2.72
10.9
Units
mg/kg
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg/day
mg/kg-day
mg/kg^day
Exposure
Route (oral,
S.C.. I.V., l.p.,
Injection)
i.p.
i.p.
oral
oral
oral
s.c.
oral
oral
Exposure
Duratlon/T
Imlnq
day 1 1 of
gestation
gestational
days 8- 14
through
day 1 5 of
the
lactation
period.
days/wk for
1 year
days/wk for
1 year
uilnj iw i
days
beginning
on days
1,7. or 13
of gestation
(22 days)
12 days
12 days
Reference
Kimbrough and
Gaines, 1968
Harbison, 1975
Deskinetal, 1979
Barnes and Denz,
1951
Barnes and Denz,
1951
Talens and Woolley,
1973
Partner et al., 1982
••
Partner el al, 1982
Comments
Toxic and teratogenic effects
were observed at this dose level.
There was a decrease in fetal
body weight and an .increase in
prenatal deaths at this dose level
Reduction in plasma
cholinesterase (pseudo ChE)
activity was the main parameter
observed, (biological
reproductive effect)
Reproductive effects were not
noted at this dose level.
Percent neonates surviving was
43%.
'Symptoms of parathion
poisoning in the dam were more
severe when parathion was
injected during the third
trimester.' Doses of 1 .5 or 2.0
mg/kg were given.
No effect on reproductive
function was- observed at this
dose level.
Egg production was reduced
during days 1 -5; cessation of
production was common between
days 6-10; other effects -
reduction in body weight, plasma
luteinizing hormone, and
proqesterone concentration.
-------
APPENDIX B Parathion - 13
Talens, G., and D. Woolley. 1973. Effects of parathion administration during gestation in
the rat on development of the young. Proc. West. Pharmacol. Soc. 16:141-145.
U.S. EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for use in Risk Assessment. P338-179874.
Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1988b. Wealth Effects Assessment for
Parathion. PBS 8-18287 8/AS. Environmental Criteria and Assessment Office, Office of
Research and Development, Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1990e. Methodology for Assessing
Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment, Washington, DC. January.
U.S. EPA (U.S. Environmental Protection Agency). 19935. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
Office of Science and Technology, Office of Water, Washington, DC.
U.S. EPA (U.S. Environmental Protection Agency). 1993c. Wildlife Exposure Factors
Handbook. Vol. I. EPA/600/R-93/187a. Office of Research and Development, Washington,
DC.
Van Leeuwen, C.J., P.S. Griffioen, W.H.A. Vergouw, and J.L. Maas-Diepeveen. 1985.
Differences in susceptibility of early life stages of rainbow trout (Salmo gairdneri) to
environmental pollutatns. Aquat. Toxicol. 7(l-2):59-78. As cited in AQUIRE (AQUatic
Toxicity_/nformation /?£trieval Database), Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency, Duluth, MN.
Van Gestel, C.A.M., E.M. Dirven-van Breemen, R. Baerselman, H.J.B. Emans, J.A.M.
Janssen, R. Postuma, and P.J.M. van Vliet. 1992. Comparison of Sublethal and Lethal
Criteria for Nine Chemicals in Standardized Toxicity Tests Using the Earthworm Eisenia
andrei. Ecotoxol. Environ. Safety 23: 206-220.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxicity - Parathion
Cas No. 56-38-2
Chemical
Name
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
Species
mallard
mallard
mallard
mallard
duckling
(MM)
sharp tailed
grouse
California
quail
Japanese
quail
pheasant
pheasant
chukar
gray
partridge
rock dove
house
sparrow
mule deer
domestic
goat
Endpolnt
mod.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
2.34
1.44
1.44
0.898
5.66
16.9
5.95
12.4
>24.0
24
16
2.52
3.36
22.0-
44.0
28.0-
56.0
Units
mg/kg-body
wl.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral,
B.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duratlon/T
Imlnq
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S! EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Comments
-
NS = Not specified
-------
Terrestrial 1. Jty - Parathion
Cas No. 56-38-2
Chemical
Name
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
parathion
Species
rat
mouse
dog
cat
rabbit
guinea pig
pigeon
chicken
quail
duck
horse
mammal
wild bird
fulvous
whistling
duck
mallard
mallard
Endpolnt
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
2
5
3
930
10
8
1330
10
4040
2100
5
49
1330
0.125-
0.250
2.4
1.9
Units
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
ug/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
ug/kg-body
wt.
mg/kg-body
wt.
ug/kg-body
wt.
ug/kg-body
wt.
mg/kg-body
wt.
ug/kg-body
wt.
ug/kg-body
wt.
mg/kg-body
wt
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral,
8.C.. I.V.. I.P..
Injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duratlon/T
Imlng
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS. 1994
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Comments
Biochemical effects.
•
•
i
-------
Freshwater Biological Uptake Measures - Parathion
Cas No. 56-38-2
Chemical
Name
parathion
parathion
parathion
parathion
Species
fish
bluegill
brook trout
brook trout
B-factor
(BCF, BAF,
BMF)
BCF
BCF
BCF
BCF
Value
52.6
462
68
344
Measured or
predicted
(m,p)
P
m
m
m
Units
Ukg
^
NS
NS
NS
Reference
Stephan, 1993
Spacie et al.. 1981 as cited in
AQUIRE, 1995
Spacie et al., 1981 as cited in
AQUIRE, 1995
Spacie et al., 1981 as cited in
AQUIRE, 1995
Comments
Normalized to 1.0% lipid.
Juvenile, 5 - 8 CM; 3 day
test.
Yearling, 60 G; 0.33 day
test.
Yearling, 60 G; 5.80 day
test.
NS = Not specified
-------
Freshwater 1 Jty - Parathion
Cas No. 56-38-2
Chemical
Name
parathion
parathion
parathion
parathion
parathion
parathion
parattiion
parathion
Species
aquatic
organisms
Daphnia
magna
Simocephalus
serrulatus
bluegill
striped bass
rainbow trout
fathead
minnow
brook trout
Endpolnt
chronic
immob.
immob.
mort.
mort.
mort.
mort.
mort.
Description
AWQC
EC50
EC50
LC50
LC50
LC50
LC50
LC50
Value
0.013
0.60 - 1 .4
(0.98)
0.37 - 0.47
(0.42)
95 - 700 (392)
17.8-2000
(328.6)
1400- 10000
(6760.8)
0.50 - 3600
(410.2)
2000
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration/
Timing
NS
48 hour
48 hour
96 hour
96 hour
96 hour
96 hour •
96 hour
Reference
51 FR 43667
Dortland, 1980 as cited in
AQUIREJ995
Sanders et al., 1966 as cited in
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
Van Leeuwen et at., 1985 as
cited in AQUIRE, 1995
AQUIRE, 1995
Spacie et al., 1981 as cited in
AQUIRE. 1995
Comments
NS = Not specified
-------
Terrestrial Biological L. *e Measures - Parathion
Cas No. 56-38-2
Chemical
Name
parathion
Species
plants
B-factor
(BCF, BAF,
BMP)
BCF
Value
28
Measured or
predicted
(m,p)
P
Units
(ug/g WW plant)/(ug/mL
soil water)
Reference
U.S. EPA. 1990E
Comments
-------
APPENDIX B Pentachlorobenzene - 1
Toxicological Profile for Selected Ecological Receptors
Pentachlorobenzene
Cas No.: 608-93-5
Summary: This profile on pentachlorobenzene summarizes the toxicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors)
for birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) arc also summarized for the ecological receptors, although some BAFs for the freshwater
ecosystem were calculated for organic constituents with log Kow between 4 and 6.5. For the
terrestrial ecosystem, these biological uptake measures also include terrestrial vertebrates and
invertebrates (e.g., earthworms). The entire toxicological data base compiled during this effort
is presented at the end of this profile. This profile represents the most current information and
may differ from the data presented in the technical support document for the Hazardous Waste
Identification Rule (HWIR): Risk Assessment for Human and Ecological Receptors
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife in
which dose-response data were reported. Developmental and reproductive toxicity studies on
laboratory mammals were limited and did not meet the minimum data set requirements of toxicity
data on at least 3 species. In spite of this, a mammalian benchmark was derived from a
subchronic toxicity study exhibiting clear dose-response data and focusing on a fetotoxic
endpoint. In this study (Khera and Villeneuve, 197.5), female Wistar rats were administered a
dietary concentration of 0, 50, 100, and 200 mg pentachlorobenzene /kg-day from gestation days
6 to 15. Khera and Villeneuve (1975) observed the fetal incidence of extra ribs and recorded a
LOAEL of 50 mg/kg-day for developmental toxicity. Under et al. (1980) conducted a subchronic
feeding study with male and female rats and observed an increase in kidney and liver weights
August 1995
-------
APPENDIX B Pentachlorobenzene - 2
and a decrease in heart weight The LOAEL in the Under et al. (1980) study was reported as
8.3 mg/kg-day.
The Khera and Villeneuve (1975) study was selected to derive the benchmark because (1) it
contains dose-response information, (2) dietary concentrations were administered via oral
ingestion during a critical life-stage period, and (3) it focused on developmental toxicity as a
critical endpoint The Under et al. (1980) study was not selected because it did not evaluate
developmental or reproductive endpoints.
The selected study LOAEL was divided by 10 to provide a LOAEL-to-NOAEL safety factor.
The LOAEL/10 from Khera and Villeneuve (1975) was then scaled for species that were
representative of the generic freshwater ecosystem using a cross-species scaling algorithm adapted
from Opresko et al. (1994)
Benchmark = NOAEL. x\
where NOAELt is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Khera and
Villeneuve (1975) study documented reproductive effects on female rats, female body weights
for each representative species were used in the scaling algorithm to obtain the lexicological
benchmarks.
Data were available on the reproductive and developmental, effects of pentachlorobenzene, as
well as chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. All of these studies identified
were conducted using laboratory rats and as such, inter-species toxicity differences were not
identifiable. Therefore, an inter-species uncertainty factor was not applied. The reproductive
LOAEL selected from Khera and Villeneuve (1975) was within an order of magnitude of the
lowest identified NEL or LEL. However, since the pentachlorobenzene data set did not meet
the minimum requirements of toxicity data on at least 3 species, the benchmarks developed for
mammals representative of an aquatic ecosystem were categorized as interim.
Birds: No subchronic or chronic studies were identified for representative of surrogate avian
species. Sources reviewed for avian toxicity information included: an on-line search of the
TOXLIT and DART databases and an extensive library search at National Institute for
Environmental Health Sciences (NIEHS) library.
Fish and aquatic invertebrates: Since a Final Chronic Value (FCV) did not exist for
pentachlorobenzene, a Secondary Chronic Value (SCV) of 1.60 mg/1 was calculated using the
Tier II methods described in Section 4.3.5. Because the benchmark for daphnids was calculated
using the Tier II method, the benchmark was categorized as interim.
August 1995
-------
m
APPENDIX B Pentachlorobenzene - 3
Aquatic Plants: The lexicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEG) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwri). Aquatic
plant data was not identified for pentachlorobenzene and, therefore, no benchmark was developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbon partition coefficient (K^) to determine a protective sediment concentration (Stephan,
1993). The EQp number is the chemical concentration that may be present in sediment while still
protecting the benthic community from the harmful effects of chemical exposure. Since a FCV
for pentachlorobenzene was not available, a Secondary Chronic Value (SCV) was calculated as
described in Section 4.3.5. The SCV was used to calculate an EQp number of 161 mg
pentachlorobenzene /kg organic carbon. Assuming a mass fraction of organic carbon for the
sediment (f^ of 0.05, the benchmark for the benthic community is 8.06 mg/kg. Since the EQp
number was based on a SCV, not an FCV, the sediment benchmark is categorized as interim.
August 1995
-------
APPENDIX B
Pentachlorobenzene • 4
Table 1. ToxicologicaJ Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
R«pr»«*niatJv«
Spotf**
mink
river otter
bald eagle
osprey
great blue heron
mallard
lessor scaup
spotted sandpiper
herring guH
kingfisher
Benchmark
Value* :
nsjjfkg-day
3.6 (i)
2.0 (i)
ID
10
ID
ID
ID
ID
ID
ID
i. SUtfy
Spodtw
rat
rat
•
-
•
•
-
-
-
-
Ettwrt
dev
dev
-
-
-
•
•
Study Vatue
Otg/Kg-day
50
50
-
,- .
-
•
-
-
Description
LOAEL
LOAEL
•
-
.
-
•
•
•
SF
10
10
-
-
• •
-
•
-
Ortgtoatswwo*
Kheraetal., 1975
Kheraetal., 1975
-
.
.
• -'
•
-
-
'Benchmark Category, a « adequate, p = provisional, i = interim; a '*' indicates that.the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B
Pentachlorobenzene - 5
Table* 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ft»pr«**fttalJv«
Spaei**
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark
Valt**
mg/L
1.6(1)
ID'
8.06(i) mg/kg
sediment
Study Sp«cJM
f
AWQC
organisms
-
AWQC
organisms
Description
scv
scv
Original Soen*
AQUIRE. 1995
-
AQUIRE. 1995
'Benchmark Category, a = adequate, p = provisional, i = interim; a '" indicates that the benchmark value was an order
of magnitude or more above the NEL or LEI for other adverse effects.
ID = Insufficient Data
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to
pentachlorobenzene. Because of the lack of additional mammalian toxicity studies, the same
surrogate-species study (Khera and Villeneuve, 1975) was used to derive the
pentachlorobenzene lexicological benchmark for mammalian species representing the
terrestrial ecosystem. The study value from Khera and Villeneuve (1975) was divided by 10
to provide a LOAEL - to - NOAEL safety factor. This value was then scaled for species
representative of a terrestrial ecosystem using a cross-species scaling algorith adapted from
Opresko et al. (1994). Since the Khera and Villeneuve (1975) study documented reproductive
effects on female rats, female body weights for each representative species were used in the
scaling algorithm to obtain the toxicological benchmarks. Since the pentachlorobenzene data
set did not meet the minimum requirements of toxicity data on at least 3 species, the
benchmarks developed for mammals representative of a terrestrial ecosystem were categorized
as interim.
August 1995
-------
APPENDIX B Pentachlorobenzene - 6
Birds: Although numerous sources were reviewed for toxicity information, no subchronic or
chronic studies were identified'for representative or surrogate avian species.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for pentachlorobenzene and, as a result,
a benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Pentachlorobenzene • 7
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
R»pt»e«m$tiV»
- Sped**
deer mouse
short-tailed shrew
meadow vole
Eastern cottontail
red fox
. raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern bobwhite
American robin
American
woodcock
plants
soil community
Benchmark
Valu."
mo/today
8.87 (i)
9.12(i)
7.41 (i)
3.13(i)
2.32 (i)
2.24 (i) -
1.12(i)
ID
ID •
ID
ID
ID
. ID
ID
Study
Specie*
rat
rat
rat
rat
rat
rat
rat
-
•-
-
-
Effect
dev
dev
dev
dev
dev
dev
dev
-
-
Origin*!
Value
mo/kfl-doy
50
50
50
50
50
50
50
•
Description
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
-
-
•
-------
APPENDIX B Pentachlorobenzene - 8
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipidrbased and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for consituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem; these models were considered appropriate to estimate BAF/s for chlordane. The
predicted BAF/1 for trophic level 4 fish in both the limnetic and littoral ecosystems is in
reasonable agreement (i.e., within a factor of 2) with the geometric mean (320,550) of the
three measured values presented in Derivation of Proposed Human Health and Wildlife
Bioaccumulation Factors for the Great Lakes Initiative (Stephan, 1993). The geometric mean
of the measured values was based on data from Oliver and Nicol (1982) and Oliver and Niimi
(1983 and 1988) for trout and salmonids. The bioconcentration factor for fish was estimated
as the geometric mean of 10 measured BCF/ values presented by Stephan (1993). Although
the predicted value of 89,474 did not differ significantly from the geometric mean of
measured values, the high quality and number of values in the data set was considered
sufficient rationale for using the geometric mean.
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of measured
values cited in Garten and Trabalka (1983). For terrestrial invertebrates, the bioconcentration
factor was estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation method is
applied to hydrophobia organic chemicals assuming that the partitioning to tissue is dominated
by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial wildlife
developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial Wildlife
from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient quality to
serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking measured
data is compared to the BBF for TCDD and that ratio (i.e., pentachlorobenzene BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. The BCF/ for earthworms was a measured value identified in a
study by Belfroid et al. (1994) on earthworm exposure to chlorobenzenes in soil. Assuming a
lipid fraction for earthworms of 0.01 (Belfroid et al., 1993), the measured value was
converted to a whole-body BCF by multiplying the lipid-based BCF/ by the the lipid fraction,
resulting in a whole-body BCF of 1.56. For hydrophobic organic constituents, the
bioconcentration factor for plants was estimated as described in Section 6.6.1 for above
August 1995
-------
APPENDIX B Pentachlorobenzene - 9
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Pentachlorobenzene • 10
Table 4. Biological Uptake Properties
MPfofliCWi
ftCtptOf
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 lish
trophic level 2
invertebrates
terrestrial
vertebrate*
terrestrial
invertebrates
earthworms
•plants
BCF, SAP, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
libU-buad or
whofe-body
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
lipid
whole-plant
valu*
213.463 (d)
202.091 (d)
177,200(1)
194,236(d)
212.985 (d)
• 439,866 (d)
62
0.0015
1,700
0.044 .
•cure*
predicted value based on
Thomann, 1989, food chain
model '
predicted value based on
Thomann, 1989, food chain
model
predicted value based on
Thomann, 1989 and adjusted to
estimate total BCF
predicted value based on
Thomann et at., 1992, food web
model
predicted value based on
Thomann et a!., 1992, food web .
model
predicted value based on
Thomann et al., 1992. food web
model
geometric mean of measured
values from Garten and
Trabalka, 1983;
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
measured value in g soi/g lipid
from Belfroid et al., 1994
U.S. EPA, 1992e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Pentachlorobenzene - 11
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
AQUIRE (AQUttiC Toxicity Information /?£trieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Banerjee, S., R.H. Sugatt, and D.P. O'Grady. 1984. A simple method for determining
bioconcentration parameters of hydrophobic cdmpounds. Envrion. Sci. Technol. 18:79-81.
As cited in Stephan, 1993. Derivations of Proposed Human Health and Wildlife
Bioaccumulation Factors for the Great Lakes Initiative, PB93-154672, Environmental
Research Laboratory, Office of Research and Development, Duluth, MN.
Barrows, M.E., S.R. Petrocelli, and K.J. Macek. 1980. Bioconcentration and Elimination of
Selected Water Pollutants by Bluegill Sunfish (Lepomis macrochirus). In: Dynamics,
Exposure and Hazard Assessment of Toxic Chemicals, R. Hague, Ed. Ann Arbor Science
Pub. Inc., Ann Arbor, MI. pp. 379-392. As cited in Stephan, 1993. Derivations of
Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative, PB93-154672, Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
Belfroid, A., A. Van Wezel, M. Sikkenk, W. Seinen, K. Van Gestel, and J. Hermens. 1994.
The toxicokinetic behavior of chlorbenzenes in earthworms (Eisenia andrei): experiments
in soil. Environmental Toxicology and Chemistry. 13:93-99.
Belfroid, A., A. Van Wezel, M. Sikkenk, K. Van Gestel, W. Seinen, and J. Hermens. 1993.
The toxicokinetic behavior of chlorbenzenes in earthworms (Eisenia andrei): experiments
in water. Ecotox. and Environ. Safety. 25:154-165.
Bruggeman, W.A., A. Opperhuizen, A. Wijenga, and O. Hutzinger. 1984. Bioaccumulation
of Super-Lipophilic Chemicals in Fish. Toxicol. Environ. Chem. 7:173-189. As cited in
Stephan, 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative, PB93-154672, Environmental Research Laboratory,
Office of Research and Development, Duluth, MN.
Carlson, A.R., and P.A. Kosian. 1987. Toxicity of chlorinated benzenes to fathead minnows
(Pimephales promelas). Arch. Environ. Contam. Toxicol. 16(2): 129-135. As cited in
Stephan, 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative, PB93-154672, Environmental Research Laboratory,
Office of Research and Development, Duluth, MN.
August 1995
-------
APPENDIX B Pentachlorobenzene - 12
Garten, C.T., and J.R: Trabalka. 1983. Evaluation of models for predicting terrestrial food
chain behavior of xenobiotics. Environ. Sci. Technol. 26(10):590-595.
Khera, K.S., and D.C. Villeneuve. 1975. Teratogenicity studies on halogenated benzenes
(Pentachloro-, pentachloronitro- and hexabromo-) in rats. Toxicol. 5:117-122.
Konemann, H., and K. Van Leeuwen. 1979. Toxicokinetics in fish: accumulation and
elimination of six chlorobenzenes by guppies. Chemosphere 9:3-19. As cited in
Stephan, 1993. Derivations of Proposed Human Health and Wildlife Bioaccwnulation
Factors for the Great Lakes Initiative, PB93-154672, Environmental Research Laboratory,
Office of Research and Development, Duluth, MN.
Under, R., T. Scotti, J. Goldstein, K. McElroy and D. Walsh. 1980. Acute and subchronic
toxicity of pentachlorobenzene. J. Environ. Pathol. Toxicol. 4:183-196
Oliver, E.G. and K.D. Nicol. 1982. Chlorobenzenes in Sediments, Water, and Selected Fish
from Lakes Superior, Huron, Erie, and Ontario. Environmental Science and Technology,
16:532:536.
Oliver, E.G., and A.J. Niimi. 1983. Bioconcentration of chlorobenzenes from water by
rainbow trout: Correlations with partition coefficients and enviromental residues.
Environ. Sci. Tech. 17:287-291.
Oliver, E.G., and A.J. Niimi. 1988. Trophodynamic analysis of polychlorinated biphenyl
congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ.
Sci. Technol. 22:388-397.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Oris, J.T., R.W. Winner, and M.V. Moore. 1991. A Four-Day Survival and Reproduction
Toxicity Test for Ceriodaphnia dubia. Environmental Toxicology and Chemistry,
10(2):217-224. As cited in AQUIRE 64Q£/atic Toxicity /nformation KEtrieval Database).
1995. Environmental Research Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, Duluth, MN.
Pereira, W.E., C.E. Rastak, C.T. Chion, T.I. Brinton, L.B, Barber, D.K. Demcheck, and C.R.
Demas. 1988. Contamination of estaurine water, biota, and sediment by halogenated
organic compunds: a field study. Environ. Sci. Technol. 22:772-778.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
August 1995
-------
APPENDIX B Pentachlorobenzene • 13
Schrap, S.M., and A. Opperhuizen. 1990. Relationship between bioavaiiability and
hydrophobicity: reduction of the uptake of organic chemicals by fish due to the sorption
on particles. Environ. Toxicol. Chem. 9:715-724. As cited in Stephan, 1993. Derivations
of Proposed Human Health and Wildlife Bioaccumulation Factors for the Great Lakes
Initiative, PB93-154672, Environmental Research Laboratory, Office of Research and
Development, Duluth, MN.
Suter n, G.W., J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-96/R1.
U.S. Department of Energy, Oak Ridge National Laboratory, Oak Ridge, TN
U.S. EPA (U.S. Environmental Protection Agency). 1985. Health Assessment Document for
Chlorinated Benzenes - Final Report. EPA/600/8-84/015F, Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1990e. Methodology for Assessing
Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment, Washington, DC. January.
U.S. EPA (U.S. Environmental Protection Agency). 1993a. Derivations of Proposed Human
Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative. PB93-
154672. Environmental Research Laboratory, Office of Research and Development,
Duluth, MN.
U.S. EPA (Environmental Protection Agency). 1993b. Technical Basis for Deriving
Sediment Quality Criteria for Nonionic Organic Contaminants for the Protection of
Benthic Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of
Water, Washington, DC.
Van Hoogen, G., and A. Opperhuizen. 1988. Toxicokinetics of chlorobenzenes in fish.
Envrion. Toxicol. Chem. 7:213-219. As cited in Stephan, 1993. Derivations of Proposed
Human Health and Wildlife Bioaccumulation Factors for the Great Lakes Initiative,
PB93-154672, Environmental Research Laboratory, Office of Research and Development,
Duluth, MN.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
Yoshioka, Y., and Y. Ose. 1993. A quantitative structure-activity relationship study and
ecotoxicological risk quotient for the protection from chemical pollution. Environ. Toxicol. Water
Quality. 8:87-101.
August 1995
-------
APPENDIX B Pentachlorophenol - 1
Toxicological Profile for Selected Ecological Receptors
Pentachlorophenol
Cas No.: 87-86-5
Summary: This profile on pentachlorophenol summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation,.and biomagnification factors)
for birds, mammals, daphnids and fish, aquatic plants and benthic organisms representing the
generic freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic
terrestrial ecosystem. Toxicological benchmarks for birds and mammals were derived for
developmental, reproductive or other effects reasonably assumed to impact population
sustainability. Benchmarks for daphnids, benthic organisms, and fish were generally adopted
from existing regulatory benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although some BAFs for the freshwater
ecosystem were calculated for organic constituents with log Kow between 4 and 6.5. For the
terrestrial ecosystem, these biological uptake measures also include terrestrial vertebrates and
invertebrates (e.g., earthworms). The entire lexicological data base compiled during this effort
is presented at the end of this profile. This profile represents the most current information and
may differ from the data presented in ihe technical support document for the Hazardous Waste
Identification Rule (HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used lo derive protective
media concentrations (CL^) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 coniains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several studies were identified which investigated subchronic and chronic effects of
pentachlorophenol exposure to mammalian species. Knudsen et al. (1974, as cited in FWS, 1989)
found thai no effecls were observed in rals fed 3 mg PCP/kg-day for 12 weeks, however, rals
given a higher dose of 50 mg PCP/kg-day exhibited kidney calcium deposils and adverse effects
on the liver. Based on ihese resulls, a NOAEL of 3 mg/kg-day and a LOAEL of 50 mg/kg-day
were inferred for toxic liver effects. The fetotoxic effects of pentachlorophenol were investigated
in rats administered oral doses containing 4, 13 and 43 mg PCP/kg-day for 181 days (Welsh et
al., 1987). Alihough no adverse effects were seen at the lowest dose, decreased fetal weights
occurred in the group given 13 mg/kg-day and embryo lethality occurred in the group maintained
at 43 mg/kg-day. Therefore, a NOAEL of 4 mg/kg-day and a LOAEL of 13 mg/kg-day was
reported for fetotoxic endpoints. Schwelz et al. (1977) conducted a study investigating the
reproductive, embryotoxic, and developmental effects of rats exposed to pentachlorophenol
August 1995
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Terrestrial Biological Uptake i. asures - Pentachlorobenzene
Cas No. 608-93-5
Chemical Name
Pentachlorobenzene
pentachlorobenzene
Pentachlorobenzene
pentachlorobenzene
Species
poultry
earthworms
earthworms
plants
B-factor
(BCF, BAF,
BMP)
BAF
BCF
BCF
BCF
Value
61.66
187000
401000
210
Measured or
predicted
(m.D)
P
m
m
P
Units
(mg/kg of
fat)/(mp/kg ot
diet)
I/kg
I/kg
(ug/gWW
plant)/(ug/mL
soil water)
Reference
Garten and Trabalka,
1983
Belfroidetal., 1993
Belfroid et al., 1993
U.S. EPA, 1990e
Comments
The worms were kept in
water, rather than in soil.
i
-------
.Freshwater Toxicity - Pentachlorobenzene
Cas No. 608-93-5
Chemical Name
Pentachlorobenzene
Pentachlorobenzene
Pentachlorobenzene
Pentachlorobenzene
Pentachlorobenzene
Species
aquatic
organisms
red killilish
Ceriodaphnia
dubia
Ceriodaphnia
dubia
Daphnia
maqna
Endoolnt
mod
mort
rep
mort-
immob
Description
LEC
LC50
'
EC50
LC50
EC50
Value
5.00E+01
2200
900-1180
(1023.3)
1100
300.4-1251.5
(446.7)
Units
ug/L
ug/L
ug/L
ug/L
uq/L
Test type
(static/ flow
through)
NA
semi static
NS
NS
NS
Exposure
Duration/
Timing
NS
48 or 96 hr;
10 yr study
96 hour
48 hour
48 hours
Reference
45 FR 79318
Yoshioka and Ose, 1993
Oris et al., 1991 as cited in
AQUIRE, 1995
Oris et al., 1991 as cited in
AQUIRE.J995
AQUIRE. 1995
Comments
The LEC is lor all
chlorinated
benzenes
1
LEC= Lowest Effects Concentrations
NS = Not specified
-------
Terrestrial Toxicity - . ^ntachlorobenzene
Cas No. 608-93-5
Chemical Name
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
-
Species
rat
rat
rat
mouse
Endpolnt
dev
hepatic, renal
mort
mort
Description
LOAEL
LOAEL
LD50
LD50
Value
50
8.3
1080
1175
Units
mg/kg-day
mg/kg-dav
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral,
S.C.. I.V.. l.p.,
Injection)
oral
oral
oral
oral
Exposure
Duration/Timing
days 6- 15 of
gestation
subchronic
NS
NS
Reference
Khera and Villeneuve,
1975
Under etal., 1980
RTECS, 1994
RTECS. 1994
Comments
Effect = fetal incidence of
extra ribs.
The following effects were
observed increased kidney
weight, a decreased heart
weight, and an increase ir
liver/body weight ratios.
NS = Not specified
-------
Freshwater Biological Uptake asures - Pentachlorobenzene
Cas No. 608-93-5
Chemical Name
pentachlorobenzene
Pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
4
Species
fish
fish
fish
fish
fish
fish
fish
fish
fish
fish
fish
fish
B-factor
(BCF, BAF.
BMP)
BAF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
6918
560
2600
708
1908
3400
3944
2607
260
2216
940
4600
Measured or
predicted
(m.p)
m
P
m
m
m
m
m
m
m
m
m
m
Units
L/kg whole-
body
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
j,
NS
Reference
Garten and Trabalka,
1983
Stephan, 1993
Konemann and van
Leeuwen, 1979 as cited
in Stephan, 1993
Barrows et al., 1980 as
cited in Stephan, 1993
Oliver and Niimi, 1983
Banerjee et al.. 1984 as
cited in Stephan, 1993
Banerjee et al.. 1984 as
cited in Stephan, 1993
Banerjee :et al.. 1984 as
cited in U.S. EPA, 1993
Bruggeman et al., 1984
as cited in Stephan,
1993
Carlson and Kosian,
1987 as cited in
Stephan, 1993
Van Hoogen and
Opperhuizen, 1988 as
cited in Stephan, 1993
Schrap and
Opperhuizen, 1990 as
cited in Stephan, 1993
Comments
Flowing water; All estimates were
calculated based on published data, the
type of studies from which the data were
taken were not specified.
Normalized to 1 .0% lipid.
Normalized to 1.0% lipid.
Normalized to 1 .0% lipid; This BCF was
based on uptake of radioactivity with no
verification of the parent chemical and
might be too high.
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
Normalized to 1.0%' lipid.
Normalized to 1 .0% lipid.
-
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
-------
Freshwater Biological Uptake Measures - Pentachlorobenzene
Cas No. 608-93-5
Chemical Name
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
pentachlorobenzene
Species
trout
salmonids
salmonids
three species
of estuarine
fish
blue crab
B-factor
(BCF, BAF,
BMP)
BAF
BAF
BSAF
BSAF
BSAF
Value
2152
6313
0.04
0.01
0.04
Measured or
predicted
(m.p)
m
m
P
P
P
Units
NS
NS
ug/g
"9/9
uo/q
Reference
Oliver and Niimi, 1983
Oliver and Niimi, 1968
Oliver and Niimi, 1988
Pereiraet. al , 1988
Pereira et. al., 1988
Comments
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
NS = Not specified
-------
APPENDIX B Pentachlorophenol - 2
dietary levels of 3 and 30 mg/kg-d. At the 30 mg/kg-d dose, there was a significant decrease
in neonatal survivability and a decrease in neonatal body weight Rats maintained on a diet
containing 3 mg PCP/kg-day exhibited no effects on reproduction or neonatal growth.
The two studies, Schwetz et al. (1977) and Welsh et al. (1987), both provide clear dose response
in establishing NOAELs based on fetotoxic effects that could impair the sustainability of a
wildlife population. The discrepancy between the two NOAELs, 3 mg/kg-d and 4 mg/kg-d, is
not significant. Therefore, the NOAEL of 4 mg/kg-day (Welsh et al., 1987) was used to
extrapolate a mammalian benchmark value because of better resolution between the NOAEL and
the LOAEL dose levels was seen in the dose regime used by Welsh. Also a technical grade 99%
PCP was used by Welsh et al. (1987) as opposed to 90.4% grade PCP used by Schwetz et al.
(1977). Although Knudsen et al. (1974, as cited in FWS, 1989) establishes a dose-response
relationship, the NOAEL was not considered suitable for deriving a benchmark value because of
uncertainty surrounding the critical endpoint Renal calcium deposits and adverse liver effects
do not clearly indicate that population sustainability would be impaired.
The NOAEL of 4 mg/kg-d from Welsh et al. (1987) was scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994)
(
Benchmark^, = NOAEL, x _ L
where NOAELt is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BWt is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Welsh et al.
(1987) study documented fetotoxic effects from pentachlorophenol exposure to male and female
mating, rats, the mean body weight of both genders of representative species was used in the
scaling algorithm to obtain the lexicological benchmarks.
Data were available on the reproductive, developmental, and growth effects of pentachlorophenol.
In addition, the data set contained studies which were conducted over chronic and subchronic
durations and during sensitive life stages. Most of the studies identified were conducted using
laboratory rats or mice, and, as such, inter-species differences among wildlife species were not
identified and, therefore an inter-species uncertainty factor was not applied. Based on the data
set for pentachlorophenol, the benchmarks developed from the Welsh et al. (1987) NOAEL of
4 mg/kg-d were categorized as adequate.
Birds: Only one study was identified which investigated developmental effects of
pentachlorophenol exposure on avian species. Prescott et al. (1982) treated 1-day old broiler
'chicks with feed containing technical grade pentachlorophenol (88% PCP) at doses of 600, 1200
or 2400 mg/kg-diet (Prescott et al., 1982). The broiler chicks were maintained on ,these dietary
levels for 8 weeks. The chickens fed the two higher doses of pentachlorophenol exhibited
;
August 1995
-------
€
APPENDIX B Pentachlorophenol - 3
decreases in body weight and liver weight, while those fed 600 mg/kg-diet showed no significant
differences from the controls in growth, histbpathology or immune response. Based on these
results, a NOAEL of 600 mg/kg-diet and a LOAEL of 1200 mg/kg-diet were inferred for
developmental effects. No information on chicken weights or consumption rates were provided
in the study. Therefore, conversion from dietary levels of pentachlorophenol in mg/kg-diet to
mg/kg-day required the use of an allometric equation:
Food consumption = 0.075CW0-8449) where W is body weight in kg (U.S. EPA, 1988).
Assuming a body weight of 1.245 kg (Prescott et al., 1985 as cited in U.S. EPA, 1988), the
NOAEL of 600 mg/kg-diet and the LOAEL of 1200 mg/kg-diet were converted to daily intakes
of 44 and 88 mg/kg-day.
The principles for allometric scaling were assumed to apply to birds, although specific studies
. supporting allometric scaling for avian species were not identified. Thus, for the aviari species
representative of a freshwater ecosystem, the NOAEL of 44 mg/kg-day (Prescott et al., 1982),
was scaled using the cross-species algorithm of Opresko et al. (1994).
For avian species, data were identified only on the developmental effects resulting from
pentachlorophenol exposure. There were no other values in the data set which were lower than
the benchmark value from Prescott et al. (1982). Laboratory experiments were not conducted
on a range of avian species and as such, inter-species differences among wildlife species were
not identifiable. Since the avian data set for pentachlorophenol did not contain the entire suite
of endpoints for population sustainability, the benchmarks developed from the Prescott et al.
(1982) study were categorized as interim.
fish and aquatic invertebrates: The Final Chronic Value (FCV) for pentachlorophenol of 1.3E-2
mg/1 was selected as the benchmark protective of daphnids and fish (U.S. EPA, 1992). The FCV
for pentachlorophenol is a pH dependent criterion calculated assuming a pH of 7.8. Since the
benchmark is based on the FCV developed for the AWQC and was slightly higher than an
identified NOEC value for rainbow trout (Dominguez and Chapman, 1984), this benchmark was
categorized as adequate*.
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEQ for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (EC^) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum). Aquatic
plant data was not identified for pentachlorophenol and, therefore, no benchmark was developed.
Benthic community. Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value (FCV) or
other chronic water quality measure, along with the fraction of organic carbon and the octanol-
carbon partition coefficient (K^) to determine a protective sediment concentration (Stephan,
1993). This methodology is applicable only for nonionic organic chemicals under the assumption
that partitioning of .the chemical between sediment-organic carbon and pore water is at
August 1995
-------
APPENDIX B Pentachlorophenol - 4
equilibrium. The ionic properties of pentachlorophenol prohibits the calculation of a benthic
community benchmark via the EQ methodology. Until a general consensus is formed on an
.appropriate methodology for deriving sediment quality values for ionic organic chemicals, the
benthic community benchmark for pentachlorophenol is under review.
August 1995
-------
APPENDIX B
Pentachldrophenol - 5
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with a Freshwater Ecosystem
ftepiwMfttailv*
Sp»oj*»
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring guN
kingfisher
'Benchmark
V.lu»
mg/kg-d
3.0 W
1.8 (a)
1.6(i)
41 (i)
37 (i)
44 (i)
49 (i)
101 (i)
45 (i)
74 (i)
Study
Specie*
rat
rat
chicken
chicken
chicken
chicken
chicken
chicken
chicken
chicken
-
Effect
feto
feto
dvp
dvp
dvp
dvp
dvp
dvp
dvp
dvp
Study Value
ma/kflH*
4
4
44
44
44
44
44
44
44
44
Owcriptien
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
.
. ,
.
.
.
Origin*
Source
Welsh et
al., 1987
Welsh et
al., 1987
Prescott et
al., 1982
Prescott et
al., 1982
Prescott et
al.. 1982
Prescott et
al., 1982
Prescott et
al., 1982
Prescott et
al., 1982
Prescott et
al., 1982
Prescott et
al., 1982
'Benchmark Category, a « adequate,
magnitude or more above the NEL or
p = provisional, i = interim; a "' indicates that the benchmark value was an order of
LEL for other adverse effects.
August 1995
-------
APPENDIX B
PenUchlorophenol • 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with a Freshwater Ecosystem
ReprwanutiV*
S|Mctft*
fish and aquatic
invertebrates
aquatic plants
benthic community
B«nchmark
.V»W
tnglL
0.013 (a*)
10
under
' review
study
Sp«cft*
aquatic
organisms
•
• '
Dwripttoe
FCV
Orfeto*
Source
U.S. EPA, 1992
-
•
IL
•Benchmark Category, a = adequate, p » provisional, i = interim; a '*' indicates that the benchmark value was an order
of magnitude or more above the NEL or LEI for other adverse effects.
ID = Insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to
pentachlorophenol. Because of the lack of additional mammalian loxicily studies, the same
surrogate-species study (Welsh et al., 1987) was used to derive the pentachlorophenol
lexicological benchmark for mammalian species representing the terrestrial ecosystem. The
study value from the Welsh et al. (1987) value was scaled for species representative of a
terrestrial ecosystem using an algorithm adapted from Opresko et al. (1994). Since the Welsh
et al. (1987) study documented fetotoxic effects from pentachlorophenol exposure to male and
female mating rats, the mean body weight of both genders of representative species was used
in the scaling algorithm to obtain the lexicological benchmarks. Based on the data set for
pentachlorophenol, the benchmarks developed from the Welsh el al. (1987) study were
categorized as adequate*.
Birds: Other than the single sludy discussed for ihe freshwater ecosystem, no additional
avian loxicily daia were identified. Therefore, ihe siudy by Prescoii el al. (1982), identifying
a developmental NOAEL of 44 mg/kg-d, was chosen 10 calculate a benchmark value for ihe
representative avian species in ihe terrestrial ecosystem. This value was iheri scaled for
species represeniative of a lerresirial ecosystem using a cross-species scaling algoriihm
August 1995
-------
APPENDIX B Pentachlorophenol - 7
adapted from Opresko et al. (1994). Since the avian data set for pentachlorophenol did not
contain the entire suite of endpoints for population sustainability, the benchmarks developed
from the Prescott et al. (1982) study were categorized as interim.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the
10th percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used.
If there were more than 10 values, the 10th percentile LOEC was used. Such LOECs applied
to reductions in plant growth, yield reductions, or other effects reasonably assumed to impair
the ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for pentachlorophenol and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Pentachiorophenol • 8
Table 3. Toxicoiogical Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Representative
Specie*
deer moose
short-tailed
shrew
meadow vole
Eastern
'cottontail
red fox
rsoooon
white- tailed
deer
red-tailed
hawk
American
kestrel
Northern
bobwhite
American
robin
American
woodcock
plant
soil community
'Benchmark
Value ms/kg-tf
8-2 (a)
8.4 (a)
7.1 (a)
2,9 (a)
2.1 (a)
2.0 (a)
1,0 (a)
45 (i)
78 (i)
71 (i)
86 (i)
72(i)
ID
10
Study
^P?^W^^
rat
rat
rat
rat
rat
rat
rat
chicken
chicken
chicken
chicken
chicken
-
-
Effect
feto
feto
feto
feto
feto
feto
feto
dvp
dvp
dvp
dvp
dvp
' :
$umy
Vaiu*
mg/kg-d
4
4
4
4
4
4
4
44
44
44
44
44
-
-
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
•
Sf -
-
-
-
•
-
-
-
•
-
-
-
-
-
• .
•„»-•.-.
Welsh et al.. 1967
Welsh et al., 1987
Welsh etal., 1987
Welsh et al., 1987
Welsh etal., 1987
Welsh et al., 1967
Welsh etal., 1987
Prescott et al.. .
1982
Prescott etal.,
1982
Prescott etal.,
1982
Prescottetal.,
1982
Prescott et al.,
1982
-
'Benchmark Category, a = adequate, p » provisional, i = interim; a '*' indicates that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Pentachlorophenol - 9
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
As stated in section 5.3.2, the BAF/s for consituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem. However, these models were considered inappropriate to estimate BAF/s for
pentachlorophenol (PCP) because they fail to consider the complex behavior of ionizable
organic compounds in surface water. The pH of the system (e.g., surface water, soil) largely
determines the extent to which the compound will dissociate into the ionized form or remain
as a neutral species. The neutral species (i.e., phenol) tends to sorb strongly to organic
carbon while the ionizable species (i.e., phenate or phenolate ions) tends to be very mobile
and is generally more significant from a toxicological standpoint. The ionization potential of
this class of organic compounds is determined by the acid dissociation constant, or pKa.
Since PCP has a pKa of approximately 4.7, PCP will be a mixture of ionized and neutral
species at environmental pH (roughly 5 to 7), dominated by neutral species at lower pH and
ionized species at higher pH. Although the Thomann (1989) model included PCP in the
comparative data set, the predicted BAF/1 value for trophic level 4 fish bf 205,838 is
substantially greater than the single measured BAF/1 of 12,589. In conjunction with the
relatively low value for BCF/1 (see below), the predicted BAF/1 appears to be unreasonably
high. The BAF/1 for trophic level 3 was estimated by multiplying the measured BAF/1 for
trophic level 4 fish by the ratio of BAF^s for trophic levels 4 and 3 predicted by the
Thomann model. While the model appears to overestimate bioaccumulation, the impact of
dissociation on trophic level 4 fish is likely to be similar in trophic level 3 fish and, therefore,
the relative bioaccumulation in each trophic level should not change appreciably. No
bioaccumulation factors were identified that were appropriate for the littoral ecosystem. The
bioconcentration factor for fish was estimated as the geometric mean of 17 measured BCF/
values presented in Stephan (1993) and the open literature. It should be noted that the
predicted value of 87,676 exceeds the geometric mean of measured values by almost a factor
August 1995
-------
APPENDIX B Pentachlorophenol • 10
of 30. The comparison of predicted and measured BCF/ values suggests that the BAF/1 of
approximately 12,000 is a more reasonable represenation of the bioaccumulation potential of
PCP.
'
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of measured.
values from Garten and Trabalka (1983). The bioconcentration factor for invertebrates was
estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation method is applied to
hydrophobic organic chemicals assuming that the partitioning to tissue is dominated by lipids.
Further, the method assumes that the BAFs and BCFs for terrestrial wildlife developed for
2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial Wildlife from TCDD and
TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient quality to serve as the standard.
The beef biotransfer factor (BBFs) for a chemical lacking measured data is compared to the
BBF for TGDD and that ratio (i.e., PCP BBF/TCDD BBF) is multiplied by the TCDD
standard for terrestrial vertebrates, invertebrates, and earthworms, respectively. The BCF/ for
earthworms was the geometric mean of measured values for several species of earthworms,
(e.g., Eisenia fetida andrei, Lumbricus rubellus, Allolobophora caliginosa) for soils of varying
compositions. For hydrophobic organic constituents, the bioconcentration factor for plants
was estimated as described in Section 6.6.1 for above ground leafy vegetables and forage
grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves and
grasses, and uptake into the plant through air diffusion. As with the aquatic ecosystem, it is
important to recognize the importance of environmental chemistry on the behavior and
bioaccumulation potential of PCP in the terrestrial ecosystem. Although measured values
were identified for vertebrates and earthworms, these values represent only a small portion of
likely environmental conditions and should be interpreted with caution. Continuing efforts
are in progress to resolve issues relevant to PCP dissociation and bioaccumulation potential in
both aquatic and terrestrial ecosystems.
August 1995
-------
APPENDIX B
Pentachlorophenol - 11
Table 4. Biological Uptake Properties
•co logical
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic-
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BMvor
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
iipid-ba*«d or
whole-body
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole- body
whole-body
whole-plant
value
1 2,589 (d)
11, 990 (d)
3,896 (t)
•
0.17
0.0014
600
0.045
•ourc*
measured value from Niimi, 1985
as cited in Thomann, 1988
trophic level 4 value adjusted by
predicted RBAF 4/3 based on
Thomann, 1 989, food chain
model
geometric mean of 17fipid-based
measured values ranging from -
200 to 70,000
insufficient data
insufficient data
insufficient data
geometric mean of measured
values in Garten and Trabalka.
1983
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCOD
geometric means of measured
values from van Gestol and
Wei-chun Ma, (1988)
U.S. EPA. 1992e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Pentachlorophenol • 12
References
Abt Associates, Inc. 1993. Revision of Assessment of risks to Terrestrial Wildlife from
TCDD and TCDF in Pulp and Paper Sludge. Prepared for Ossi Meyn, U.S.
Environmental Protection Agency, Office of Pollution Prevention and Toxics.
Adema, D.M.M. and G.J. Vink. 1981. A Comparative Study of the Toxicity of 1,1,2-
Trichloroethane, Dieldrin, Pentachlorophenol and 3,4-Dichloroaniline for Marine and Fresh
Water Organisms, Chemosphere, Vol. 10, No. 6, pp. 533-554.
AQUIRE (AQUatic Toxicity Information REtrieval IDatabase). 1995. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Belfroid, A., A. Van Wezel, M. Sikkenk, W. Seinen, K. Van Gestel, and J. Hermens. 1994.
The toxicokinetic behavior of chlorbenzenes in earthworms (Eisenia andref): experiments
in soil. Environmental Toxicology and Chemistry. 13:93-99.
Belfroid, A., A. Van Wezel, M. Sikkenk, K. Van Gestel, W. Seinen, and J. Hermens. 1993.
The toxicokinetic behavior of chlorbenzenes in earthworms (Eisenia andref): experiments
in water. Ecotox. and Environ. Safety. 25:154-165.
Choudhury, H., J. Coleman, C. T. De Rosa, and J. F. Stara. 1986. Pentachlorophenol:
health and environmental effects profile. Toxicol. Ind. Health 2:483-571.
Courtney, K.D., M.F. Copeland, and A. Robbins. 1976. The Effects of
Pentachloronitrobenzene, Hexachlorobenzene, and Related Compounds on Fetal
Development. Toxicology and Applied Pharmacology, 35:239-256.
Dominguez, S. E., and G. A. Chapman. 1984. Effect of pentachlorophenol on the growth
and mortality of embryonic and juvenile steelhead trout. Archives of Environmental
Contamination and Toxicology 13:739-743. ,
Freitag, D., H. Geger, A. Kraus, R. Viswanathan, D. Kotzias, A. Assar, W. Klein, and
F. Korte. 1982. Ecotoxicological profile analysis. VII. Screening chemicals for their
environmental behavior by comparative evaluation. Ecotoxicol. Environ. Saf. 6:60-81.
As cited in Agency for Toxic Substances and Disease Registry (ATSDR). 1987.
Toxicological Profile for Pentachlorophenol. Public Health Service, U.S. Department of
Health and Human Services, Atlanta, GA.
Garten, Jr., C. T., and J. R. Trabalka. 1983. Evaluation of models for predicting terrestrial
food chain behavior of xenobiotics. Environmental Science and Technology 17:590-595.
August 1995
-------
APPENDIX B Pentachlorophenol - 13
Haimi, J., J. Salmineh, V. Huhta, J. Knuutinen, and H. Palm. 1992. Bioaccumulation of
Organochlorine Compounds in Earthworms. Soil Biol. Biochem., Vol. 24, No. 12, pp.
. 1699-1703.
Hattula, M. L., V. M. Wasenius, H. Reunanen, and A. U. Arstila. 1981. Acute toxicity of
some chlorinated phenols, catechols and cresols to trout. Bull. Environ, Contam. Toxicol.
26:295-298.
Haque, A., and W. Ebing. 1988. Uptake and accumulation of pentachlorophenol and sodium
pentachlorophenate by earthworms from water and soil. The Science of the Total
Environment 68:113-125.
Hedtke, S. FM C. W. West, K. N. Allen, T. J. Norberg-King, and D. I. Mount 1986.
Toxicity of pentachlorophenol to aquatic organisms under naturally varying and controlled
environmental conditions. Environ. Toxicol. Chem. 5(6):531-542.
Hill, E.F. and M.B. Camardese. 1986. Lethal dietary toxicities of environmental
contaminants and pesticides to Cotumix. U.S. Fish Widl. Serv. Fish Wildl. Tech. Rep. 2.
147 pp. As cited in Eisler, R. 1989. Pentachlorophenol Hazards to Fish, Wildlife, and
Invertebrates: A Synoptic Review. U.S. Fish Wildl. Serv. Bioli Rep. 85 (1.17). 72 pp.
Holcombe, G. W., G. L. Phipps, and J. T. Fiandt. 1982. Effects of phenol, 2,-4
dimethylphenol, 2,4-dichlorophenol, and pentachlorophenol on embryo, larval, and early-
juvenile fathead minnows (Pimephales promelas). Archives of Environmental
Contamination and Toxicology 11:73-78.
Inglis, A., and E. L. Davis. 1972. Effects of water hardness on the toxicity of several
organic and inorganic herbicides to fish. Bur. Sport Fish. Wildl. Tech. Paper No. 67
U.S.D.I. 22 pp.
Kobayashi, K., and H. Akitake. 1975. Studies on the metabolism of chlorophenols in fish.
I. Absorption and excretion of PCP by goldfish. Bull. Jpn. Soc. Sci. Fish 41:87-92.
Knudsen, I., H. G. Verschuuren, E. M. D. Tonkelaar, R. Kroes, and P. F. W. Helleman.
1974. Short-term toxicity of pentachlorophenol in rats. Toxicology 2:141-152.
Loehr, R. C., and R. Krishnamoorthy. 1988. Terrestrial bioaccumulation potential of
phenolic compounds. Hazardous Waste and Hazardous Materials Vol. 5, No. 2.
Makela, T. P., T. Pentanen, J. Kukkonen, and A. O. J. Oikari. 1991. Accumulation and
depuration of chlorinated phenolics in the freshwater mussel (Anodonta anatina L.).
Ecotoxicology and Environmental Safety 22:153-163.
August 1995
-------
APPENDIX B Pentachlorophenol - 14
Niimi, A. J., and L. A. McFadden. 1982. Uptake of sodium pentachlorophenate (NaPCP)
from water by rainbow trout (Salmo gairdneri) exposed to concentrations in the ng/litre
range. Bulletin of Environmental Contamination & Toxicology 28.(1):11-19.
Prescott, C. A., B. N. Wilke, B. Hunter, and R. J. Julian. 1982. Influence of a purified grade
of pentachlorophenol on the immune response of chickens. Am. J. Vet. Res. 43:481-487.
I ,
Pruitt, G. W., B. J. Grantham, and R. H. Pierce. 1977. Accumulation and elimination of
pentachlorophenol by the bluegill Lepomis macrochirus. Am. Fish. Soc. 106:462-465.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health, Washington, DC.
Schwetz, B.A., P.A. Keeler, P.J. Gehring. 1974. The effect of purified and commercial grade
pentachlorophenol on rat embryonal and fetal development. Toxicology and Applied
Pharmacology, 28:151-161.
Schwetz, B. A., J. F. Quasi, P. A. Keeler, C. G. Humiston, and R. J. Kociba. 1978. Results
of two-year toxicity and reproduction studies on pentachlorophenol in rats. pp. 301-309.
In K. R. Rao (ed.). Pentachlorophenol: Chemistry, Pharmacology, and Environmental
Toxicology. Plenum Press, New York.
Smith, A. D., A. Bharath, C. Mallard, D. Orr, L. S. McCarty, and G. W. Ozburn. 1990.
Bioconcentration kinetics of some chlorinated benzenes and chlorinated phenols in
American flagfish, Jordanella floridae (Goode and Bean). Chemosphere 20:379-386.
Stephan, C.E. 1993. Derivations of proposed human health and wildlife bioaccumulation
factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
U.S. EPA (U.S. Environmental Protection Agency). 1980. Ambient Water Quality for
Pentachlorophenol. U.S. Environmental Protection Agency Rep. 440/5-80-065. 89 pp.
As cited in Eisler, R. 1989. Pentachlorophenol Hazards to Fish, Wildlife, and
Invertebrates: A Synoptic Review. U.S. Fish Wildl. Serv. Biol. Rep. 85 (1.17). 72 pp.
U.S.EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. PB88-179874.
Environmental Criteria and Assessment Office, Office of Health and Environmental
Assessment, Office of Research and Development, EPA/600/6-87/008.
U.S. EPA (Environmental Protection Agency). 1990e. Methodology for Assessing Health
Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office
of Health and Environmental Assessment, Washington, DC. January.
August 1995
-------
APPENDIX B Pentachlorophenol - 15
U.S. EPA (Environmental Protection Agency). December 22, 1992b. Water quality
standards; establishment of numeric criteria for priority toxic pollutants; State's
compliance. Federal Register 57(No. 246):60912.
van Gestel, C. A. M., and W. Ma. 1988. Toxicity and bioaccumulation of chlorophenols in
earthworms in relation to bioavailability in soil. Ecotoxicology and Environmental Safety
15:289-297.
i
Veith, G. D., D. L. DeFoe, and B. V. Bergstedt. 1979b. Measuring and estimating the
bioconcentration factor of chemicals in fish. Journal of the Fisheries Research Board of
Canada 36:1040-1048.
Welsh, J. J., T. F. X. Collins, T. N. Black, et al. 1987. Teratogenic potential of purified
pentachlorophenol and pentachloroanisole in subchronically exposed Sprague-Dawley rats.
Food and Cosmetics Toxicology 25:163-172.
Will, M.E. and G.W. Suter, 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial Toxiclty jntachlorophenol
Cas No. 87-86-5
Chemical Name
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
Species
mouse
mouse
rat
rat
rat
rat
rat
female rat
male rat
rat
Endpoint
let
let
fet
emb.fet
rep, dvp,
emb
liver, kidney
liver, kidney
liver, kidney
Description
NOEL
NOEL
AEL
LOAEL
NOAEL
LOEL
NOEL
NOEL
NOEL
NOAEL
Value
3
10
75
13
4
5
3
3
.10
3
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
s.c., i.v., i.p.,
injection)
oral
oral
oral
oral
oral
gavage
oral
oral
oral
oral
Exposure
Duration
/Timing
24 months
22 months
gestation days
7-18
181 days
181 days
days 6- 15 of
gestation
62 d prior to
mating + 15d
during mating
+ until 21 d of
weaning
24 months
22 months
1 2 weeks
Reference
EPA 1980 as cited in Eisler,
1989
EPA 1980 as cited in Eisler.
1989
Courtney et al., 1976
Welsh el al., 1987
Welsh etal:, 1987
Schwetz et al., 1974
Schwetz et al., 1977
Schwetz et al., 1977
Schwetz et al., 1977
Knudsenetal.. 1974
Comments
No measurable effect in females.
No measureable effect in males.
sign, decreased mean fetal body
weight.
Decreased fetal body weight.
Purified pep was 99% pep. Doses
were 0,5,15,30,50 mg/kg-d. at 5
mg/kg-d cranial ossification was
delayed sign.
pep sample was 90.4 % pep. Doses
were 0,3,30 mg/kg-d. At 30 dose
there was sign, decrease in % of
livebom pups. Other dev. effects at
30 mg/kg-d during weaning period.
doses were 1.3,10,30 mg/kg-d. At
30 and 10 mg/kg-d there was an
accumulation of pigments in liver
and kidneys.
doses were 1,3,10,30 mg/kg-d. At
30 mg/kg-d there was an
accumulation of pigments in liver
and kidneys.
No observable effects.
-------
Terrestrial Toxicity - Pentachlorophenol
Cas No. 87-86-5
Chemical Name .
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
Species
rat
doq
guinea pig
chicken
chicken
mallard
pheasant
Japanese
quail
Endpoint
iver, kidney
acute
acute
%
dvp
dvp
acute
acute
acute
Description
LOAEL
LD50
LD50
LOAEL
NOAEL
LD50
LD50
LD50
Value
50
150-200
100
87
43.5
380
504
5.139
Units
mg/kg-day
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-diet
mg/kg-d
mg/kg
mg/kg
mq/kq
Exposure
Route (oral,
s.c., i.v., i.p.,
injection)
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
12 weeks
NS
NS
8 weeks
8 weeks
NS
NS
5 days
Reference
Knudsenetal., 1974
Knudsenetal., 1974
Choudhury et al., 1986
Prescott et al., 1982
-
Prescott et al., 1982
Hudson etal., 1970
Hudson etal., 1970
Hill and Camardese, 1986 as
cited in Eisler. 1989
Comments
Adverse effects on liver, kidney
calcium deposits and blood
chemistry.
. ,
Doses were 0,600,1200,2400. Body
weight decreased, dose calculated
from 1200 ppm, body wt.=1 .245 kg
(Prescott et al., 1982 & 1985 as
cited in U.S.EPA.1988), Food
intake=0.09025 kg/d (U.S.EPA,
1988)
No significant difference from •
controls in growth, blood chemistry,
histopathology or immune
response. Dose calculated from 600
ppm, body wt.=1 .245 kg (Prescott e
al., 1982 & 1985 as cited in
U.S.EPA.1988), Food
intake=0.09025 kg/d (U.S.EPA,
1988)
birds age 14 days were fed treated
diets for 5 days, then untreated feec
for 3 days
NS = Not specified
i
-------
Freshwater Toxicitv entachlorophenol
Cas No. 87-86^5
Chemical Name
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
penlachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
Species
rainbow trout
rainbow trout
fathead
minnow
Sheepshead
minnow
Daphnia
magna
channel
cattish
bluegill
rainbow
trout
fathead
minnow
aquatic
organisms
Endpoint
chronic
rep, dvp
chronic
chronic
acute
acute
acute
acute
acute
chronic
Description
NOEC
LOEC
NOEC
NOEC
LC50
LC50
LC50
LC50
LC50
NAWQC
Value
11
19
45
47
55-2790
(736)
54-132 (73)
24-270(143)
18-3,000
(253)
95-600 (243)
0.013
Units
ug/L
ug/L
ug/L
ug/L
ug/l
ug/I
ug/L
ug/L
ug/L
mo/L
. Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
from
fertilization
through day
72
from
fertilization
through day
72
lifetime
exposure
lifetime
exposure
2 day
2 day
NS
NS
NS
NS
Reference
Dominguez and Chapman,
1984
Dominguez and Chapman,
1984
EPA, 1980 as cited in Eisler.
1989; Holcombe el al., 1982
EPA, 1980 as cited in Eisler,
1989
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
U.S. EPA. 1992b
Comments
Negligible embryonic
mortality, alevin mortality 3)
greater than control; alevin
growth affects.
NS = Not specified
-------
Freshwater Biological Uptake Measures - Pentachlorophenol
Cos No. 87-86-5
Chemical Name
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
Species
rainbow trout
rainbow trout
freshwater mussel
goldfish
(lathead minnow
golden orfe
rainbow trout
rainbow trout
fish
fish
fish
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BAF
BAF
Value
286x-572x
/
160x
81-263
56.00
776.00
1,047.00
251
5,370.00
11
776
129
Measured
or
Predicted
(m,P)
NS
NS
m
NS
NS
NS
NS
NS
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
L/kg
L/kg
Ukq .
Reference
Choudhury et al., 1986; Niimi
and McFadden, 1982
Choudhury et al., 1986; Niimi
and McFadden, 1982
Makelaetal . 1991
Loehr and Krishnamoorthy,
1988
Loehr and Krishnamoorthy,
1988
Loehr and Krishnamoorthy,
1988
Loehr and Krishnamoorthy,
1988
Loehr and Krishnamoorthy,
1988
U.S. EPA, 1992
Garten and Trabalka, 1983
Garten and Trabalka. 1 983
Comments
After 115 days at 0.035 ug PCP/L
of medium.
After 115 days at 0.7 ug PCP/L of
medium.
Normalized to 3% lipid.
Flowing water; All estimates were
calculated based on published
data, the type of studies from
which the data were taken were
not specified.
Microcosm; All estimates were
calculated based on published
data, the type of studies from
which the data were taken were
not specified.
-------
Freshwater Biological Uptake .easures - Pentachlordphenol
Cos No. 87-86-5
Chemical Name
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
Species
fish
fish
fish
Carassius auratus
Lepomis macrochirus
Salmo trutta
Salmo gairdneri
Salmo gairdneri
Leuciscus idus
melanotus
B-factor
(BCF, BAF.
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
560.00
101.30
16.30
1,000.00
320.00
200.00
600.00
232.00
1,050.00
Measured
or
Predicted
(m,p)
P
m
m
m
m
m
m
m
m
Units '
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Stephen, 1993
Veithelal., 1979
Smith etal., 1991
Kobayashl and Akitake,1975
Pruitt etal., 1977
Hartulaelal , 1981"
Niimi and McFadden, 1982
Niimi and McFadden, 1982
Freitag et al , 1982 as cited in
ATSDR. 1987
Comments
Normalized to 1% lipid.
Normalized to 1% lipid.
Normalized to 1% lipid.
5 days exposure to 100 ug/L.
1 day exposure to 100 ug/L.
1 day exposure to 200 ug/L.
65 day exposure to .035 ug/L.
65 day exposure to .660 ug/L.
NS = Not specified
-------
Terrestrial Biojogical Uptake Measures - Pentachlorophenol
Cos No. 87-86-5
Chemical Name
pentachlorophenol
Pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
pentachlorophenol
Species
poultry
earthworms
earthworms
earthworms
(Eisenia tetida
andrei)
earthworms
(Eisenia fetida
andrei)
earthworms
(Lumbricus
rubellus)
earthworms
(Lumbricus
rubellus)
plant
B-tactor
(BCF, BAF.
BMP)
BAF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Value
0.17
2.9
2.4
5.3
3.4
4
8
0.045
Measured
or
Predicted
(m,p)
NS
m
m
m
m
m
m
p
units
(mg/kg of
fat)/(mg/kg of diet)
NS
NS
NS
NS
NS
NS
(ug/g DW
plant)/(ufl/g soil)
Reference
Garten and Trabalka, 1983
Hague and Ebinq, 1988
Haque and Ebing, 1988
van Gestel and Wei-chun Ma,
1988
van Gestel and Wei-chun Ma,
1988
van Gestel and Wei-chun Ma,
1988
van Gestel and Wei-chun Ma,
1988
U.S. EPA, I990e
Comments
Uptake by earthworms from aqueous
solution; 1 mg/L
Uptake by earthworms from aqueous
solution; 10 mg/L
Uptake by earthworms from Kooyenburg
soil (very humic sand)
Uptake by earthworms from Molten soil
(moderately humic sand)
Uptake by earthworms from Kooyenburg
soil (very humic sand)
Uptake by earthworms from Holten soil
(moderately humic sand)
NS = Not specified
-------
APPENDIX B Polychlorinated Biphenyl (PBC) - Arocldr - 1
lexicological Profile for Selected Ecological Receptors
Polychlorinated Biphenyl (PCB) - Aroclor 1254
CasNo.: 11097-69-1
Summary: This profile on polychlorinated biphenyls (PCBs) summarizes the lexicological
benchmarks and biological uptake measures (i.e., bioconcentration, bioaccumulation, and
biomagnification factors) for birds, mammals, daphnids and fish, aquatic plants and benthic
organisms representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. The toxicily data presented in this profile
frequently refer to Aroclor 1254 because: (1) Aroclor 1254 is one of the most lexicologically
potenl PCB mixlures, (2) toxicity data on Aroclor 1254 was available on most ecological
receplors, and (3) Aroclor 1254 was one of the higher volume PCB mixlures produced. While
recognizing lhal basing ihe ecological benchmarks on Aroclor 1254 may be conservative, there
are currently no available melhods to select an individual PCB congener or mixture to represent
tolal PCBs. Consequentiy, il was determined lhal daia availability on a high volume mixture was
an appropriate approach for ihe development of ecological benchmarks for total PCBs.
Toxicological benchmarks for birds and mammals were derived for developmental, reproductive
or other effects reasonably assumed to impaci population susiainabilily. Benchmarks for
daphnids, benihic organisms, and fish were generally adopled from existing regulaiory
benchmarks (i.e., Ambieni Water Qualily Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constiiuenis wilh log Kow beiween 4 and 6.5. For ihe terrestrial ecosystem,
ihese biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological daia base compiled during this effort is presented ai ihe
end of ihis profile. This profile represenls ihe mosi currenl information and may differ from the
data presented in the technical support document for ihe Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents ihe rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for ihe generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 coniains
benchmarks for aquatic organisms in the limnetic and litioral ecosystems, including aquatic
planis, fish, invertebrates and benthic drganisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Several loxicily sludies were identified that focused on the effects of Aroclor-1254
on laboratory animals, or explored ihe effecis of oiher PCB congeners on wildlife and laboratory
August 1995
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APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 2
animals. Numerous wildlife studies have demonstrated that mink are among the most sensitive
mammalian species to the toxic effects of PCBs, with some PCB congeners being more toxic
than others (U.S. EPA, 1993b). The main chronic effects documented in minks as a result of
dietary exposure to PCBs, have been decreased reproductive success, as evidenced by reduced
whelping rates, fetal death, and reduced growth among the young.
Truelove et al. (1982) dosed three cynomolgus monkeys with 100 or 400 ug/kg-day of Aroclor-
1254 from approximately 60 days of gestation. The two monkeys dosed with 100 ug/kg-day
delivered stillborn infants and the 400 ug/kg-day dosed monkey delivered a term infant that had
impaired.irnmunologic function compared with the control infant Platonow and Karstad (1973)
administered Aroclor-1254 to Jersey cows, and fed the resulting contaminated beef to mink over
160 days at 0.64 and 3.57 ppm total PCBs. At 0.64 ppm total PCBs in the diet, 2 of 14 adult
mink died before the end of the experiment and only 1 of 12 mink produced kits. The three kits
produced died during the first day after birth. Hornshaw et al. (1983) fed Great Lakes fish or
fish products to mink for up to 290 days. Dietary concentrations of PCB residues ranged from
0.21 to 1.50 ppm. Only mink fed PCBs at concentrations of 0.21 ppm had reproduction and kit
survival similar to the controls. Mink fed a diet containing 0.48 ppm PCB residues had inferior
reproductive performance and/or kit survival when compared -to the controls. Aulerich and
Ringer (1977) exposed mink to dietary Aroclor-1254 at 0, 5, and 10 ppm over a 9-month period.
All of the mink fed PCB-supplemented diets failed to produce offspring. In a subsequent
experiment, mink were provided diets containing 2 ppm Aroclor 1016, 1221, 1242, or 1254, and
monitored over 297 days. Aroclor 1254 was the only PCB that had an adverse effect on
reproduction at the 2 ppm dietary level. Only 2 of 7 females whelped and produced only 1 live
kit which weighed considerably less than the average weight of the kits whelped by the females
on the other dietary treatments. In additional research, Aulerich and Ringer (1977) documented
the reproductive sensitivity of the female mink to 4-month dietary doses of 1, 5, and 15 ppm of
Aroclor-1254. At the 5 ppm and the 15 ppm levels, the female minks exhibited markedly
reduced reproduction, however, reproduction rates, at the 1 ppm level were not significantly
inhibited.
The Truelove et al. study (1982) was not chosen for the development of a wildlife benchmark
for mammals because the monkey is taxonomically further from the representative aquatic
mammals when compared to the mink (Aulerich and Ringer study 1977). Also, the low number
of subjects and doses used in the Truelove (1982) study, limits the confidence in the dose-
response correlation. According to Platonow and Karstad (1973) and Hornshaw et al. (1983),
reproductive impairment occurs in mink at even lower concentrations when the PCBs fed to the
mink have first been metabolized by another species. However, these studies are not appropriate
for the development of a wildlife benchmark value because possible contamination of feed by ,
additional environmental contaminants was not investigated. Therefore, the Aulerich and Ringer
(1977) study, which had a sufficient dose range and documented toxic effects specific to Aroclor-
1254, was used to extrapolate a benchmark value for aquatic mammals. Using the average
female mink body weight in the study (0.974 kg) and a daily food intake rate of 0.11 kg/d (U.S.
EPA, 1993a), the NOAEL calculated for reproductive effects to mink was 0.12 mg/kg-d
(equivalent to a 1 ppm dietary dose).
August 1995
-------
APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 3
The study value from the Aulerich and Ringer (1977) was scaled for species representative of
a freshwater ecosystem using the cross-species scaling algorithm adapted from Opresko et al.
(1994)
Benchmark^ = NOAEL, x - L
where NOAELt is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Aulerich and
Ringer (1977) study documented the sensitivity of the reproductive physiology of the female
mink to Aroclor- 1254 (e.g., no apparent adverse effects on male spermatogenesis), female body
weights for each representative species were used in the scaling algorithm to obtain the
toxicological benchmarks.
Data were available on the reproductive, developmental, and growth effects of Aroclor- 1254
exposure. In addition, the data set contained studies which were conducted over chronic and
subchronic durations and during sensitive life stages. The study value selected from the Aulerich
and Ringer (1977) was a NOAEL based on a developmental endpoint that was within an order
of magnitude of the lowest identified NEL or LEL. An interspecies uncertainty factor to account
for differences in toxicological sensitivity was not supported by the data set Based on the data
set for Aroclor-1254, the benchmarks developed from the Aulerich and Ringer (1977) study were
categorized as adequate.
Birds: Chronic toxicity studies have, been conducted on mallards, Japanese quail, pheasants, and
domestic leghorn chickens. In a subchronic study, Platnow and Reinhart (1973) exposed chickens
to dietary concentrations of 0, 5 or 50 ppm Aroclor-1254 for up to 39 weeks. A significant
decline in production and hatchability of fertile eggs was observed among hens maintained at the
50 ppm level. At 5 ppm, egg production was reduced, but not the hatchability of the fertile eggs.
After the first 14 weeks of exposure, Platnow and Reinhart (1973) noted a significant decline in
the fertility of the 5 ppm group (the 5 ppm dose corresponded to a daily dose of 2.44 mg/kg-d).
In another study, Lillie et al. (.1974) assessed the reproductive effects of dietary exposure to either
2 or 20 ppm Aroclor-1254 on chickens. Reduced egg production and egg hatchability were
observed only among the group of chickens maintained on 20 ppm Aroclor-1254. In this study,
Lillie et al. (1974) also monitored the growth and survival of chicks produced from hens exposed
to Aroclor-treated feed. A significant reduction in growth was observed among chicks produced
from hens maintained on feed treated with Aroclor-1254 at 2.0 and 20 ppm. The NOAEL of 2
ppm in the Lillie et al. (1974) study corresponded to a daily dose of 0.98 mg/kg-d. Dahlgren et
al. (1972) assessed the reproductive effects of orally-administered Aroclor-1254 on the ring-
necked pheasant. Female pheasants were dosed once per week, via gelatin capsule, at rates of
0, 12.5, and 50 mg/week, and male pheasants at rates of 0 and 25 mg/week, for 16 weeks. Egg
production and chick survivability were significantly reduced among hens administered 50 mg
Aroclor-1254 per week, but not among hens administered 12.5 mg per week. Although no effect
August 1995
-------
APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 4
of exposure to Aroclor-1254 on egg fertility was noted, significant reductions in hatchability were
reported among eggs from both hen treatment groups. Using a pheasant body weight of 1 kg
(U.S.EPA, 1993b), the LOAEL of 12.5 ppm for egg hatchability was converted to a daily dose
of 1.8 mg/kg-d.
The pheasant study by Dahlgren et al. (1972) was used to derive the avian benchmark value for
the freshwater ecosystem. Pheasants have been shown to be as sensitive to PCB exposure as
laboratory chickens and the toxic endpoint of egg hatchability is a meaningful reproductive effect
associated with avian dietary exposure to PCBs. In addition to these reasons, the Dahlgren et
al. (1972) study was chosen over the Lillie et al. (1974) study, because pheasants, more so than
laboratory chickens, are considered a wildlife species that may have taxonomic and habitat
similarities with the representative avian species in Table 1.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. The. LOAEL of 1.8 mg/kg-d
value from the Dahlgren et al. (1972) was divided by 10 to provide a LOAEL-to-NOAEL safety
factor. The value was then scaled using the cross-species scaling method of Opresko et al.
(1994).
Data were available on the reproductive and developmental effects of Aroclor-1254, as well as
on growth or survival. In addition, the data set contained studies that were conducted over
chronic and subchronic durations. There were no other values in the data set which were more
than a magnitude lower than the benchmark value. Laboratory experiments of similar types were
not conducted on a range of avian species and as such, inter-species differences among wildlife
species were not identifiable. Based on the avian data set for toxaphene, the benchmarks
developed from the Dahlgren et al. (1972) study were categorized as provisional.
Fish and aquatic invertebrates: Since an AWQC was not available for Aroclor-1254, the Tier
II methodology described in Section 4.3.5 was used to calculate a Secondary Chronic Value
(SCV) for Aroclor-1254 of 1.9E-4 mg/1. Suter and Mabrey (1994) calculated an SGV of 2.0E-05,
however, the data set did not include a daphnid value nor contain as many data points as the
SCV calculated from AQUIRE. Tier II values are developed so that aquatic benchmarks could
be derived for chemicals lacking the necessary data to calculate an FCV. The SCV of 1.9E-4
mg/1 was selected as the benchmark protective of daphnids, fish, and other aquatic organisms.
The benchmark was categorized as interim, since its basis was a SCV calculated from AQUIRE.
Aquatic Plants: The lexicological benchmarks for aquatic, plants were either (1) a no observed
effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular
aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of
freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwn). The
aquatic plant benchmark for Aroclor-1254 is 1E-4 mg/L based on a reduction in carbon fixation
by Scenedesmus quadricaudata (Suter and Mabrey, 1994). As described in Section 4.3.6, all
benchmarks were described as interim.
i
Benthic community: Benchmarks for the protection of benthic organisms were determined using
the Equilibrium Partition (EQp) method. The EQP method uses a Final Chronic Value (FCV) or
August 1995
-------
1
APPENDIX B
Polychlorinated Biphenyl (PBC) • Aroclor - 5
SCV, along with the fraction of organic carbon and the octanol-carbon partition coefficient
to determine protective sediment concentration (Stephan, 1993). The EQp number is the chemical
concentration that may be present in the sediment while still protecting the benthic community
from harmful effects from chemical exposure. The SCV, calculated from the AQUIRE database,
for Aroclor-1254 was used to calculate an EQp value of 290 mg Aroclor-1254 /kg organic
carbon. Assuming a mass fraction of organic carbon for the sediment (f^ of 0.05, the
benchmark for the benthic community is 14.5 mg/kg sediment. Since the EQp number was based
on an SCV, the sediment benchmark was categorized as interim.
Table 1. Tpxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
JtapiMMUtiw
)p*ciM
mink
river otter
bald eagle
osprey
great blue
heron
mallard
lesser scaup
spotted
sandpiper
herring gull
kingfisher
- fendmw*
YfKWmvfcgKi
0.1 2 (a)
0.07 (a)
0.1 2 (p)
0.16(p)
0.1S(p)
0.18 (p) ,
0.19(p)
0.39 (p)
0.18(p)
0-29 (p)
Study
•frurix
mink
mink
pheasant
pheasant
pheasant
pheasant
pheasant
pheasant
pheasant
pheasant
en**
rep
rep
rep
rep
rep
rep
rep
rep
rep
rep
, *««r
V»k*
«»9*(H^
0.12
0.12
1.8
18
1.8
1.8
1.8
1.8
1.8
1.8
OMcriprtnn
NOAEL
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
»
•
10
10
10
10
10
10
10
10
<*I|frlt»MMM»
Autorich and Ringer,
1977
Aulerich and Ringer,
1977
Dahlgrer et al.. 1972
Dahlgran et al., 1972
Dahlgren et al.. 1972
Oahlgren et al., 1972
Dahlgren et al., 1972
Oahlgren et al., 1972
Dahlgren et al., 1972
Dahlgren et al., 1972
•Benchmark Category, a = adequate, p » provisional, i = interim; a "" indicates that the benchmark value was an order of magnitude
or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Polychlorinated Biphenyl (PBC) - Aroclor - 6
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Aquatic Ecosystem
feptM*nitttv«
8p*ei«*
fish and aquatic
invertebrates
aquatic plants
benthic
community
Benchmark.
Vitu.
*tft
1 .9E-4 (i )
1.0E-4(i)
14.5 (i)
mg/kg
sediment
• Study Sf»cfe«
aquatic
organisms
aquatic plants
benthic
community
*^
scv
cv
SCV x Koc
OriofcatSouN*
AQUIRE. 1995
SutBf 4
Mabrey, 1994
AQUIRE, 1995
'Benchmark Category, a = adequate, p = provisional, i = interim; a'" indicates that the benchmark value was an order of magnitude
or more above the NEL or LEL for other adverse effects.
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial Enq&m
This section presents the rationale behind lexicological benchmarks used to derive protective media
concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains benchmarks for mammals,
birds, plants, and soil invertebrates representing the generic terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, several toxicity studies
were identified that focused on the effects of Aroclor-1254 on mink. Since no additional studies for
terrestrial mammals were found, the same, surrogate study (Aulerich and Ringer, 1977) was used to
calculate benchmark values for mammalian species representating the general terrestrial ecosystem. The
NOAEL from the Aulerich and Ringer et al. (1977) study was scaled for species in the terrestrial
ecosystem using the cross-species scaling algorithm adapted from Opresko et al. (1994). Since the
Aulerich and Ringer et al. (1977) study documented reproductive effects from Aroclor-1254 exposure
to female minks, female body weights for each representative species were used in the scaling algorithm
to obtain the toxicological benchmarks. Based on the data set for Aroclor-1254, the benchmarks
developed from the Aulerich and Ringer et al. (1979) study were categorized as adequate.
Birds: No additional avian toxicity studies were identified for species representing the terrestrial
ecosystem. Thus, for avian species in the terrestrial ecosystem, the LOAEL/10 of 0.18 mg/kg-day from
the Bush et al. (1977) study was used as the benchmark value. This value was then scaled for a
terrestrial species using the cross-species scaling algorithm adapted from Opresko et al. (1994). Based
on the avian data set for toxaphene, the benchmarks developed from the Dahlgren et al. (1972) study
were categorized as provisional.
August 1995
-------
APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 7
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root length. Phytotoxicity benchmarks were selected as the lowest concentration identified for
plant growth, yield reductions, or other effects reasonably assumed to impair the ability of a plant
population to sustain itself, such as a reduction in seed elongation. The benchmark for terrestrial plants
was 40 mg/kg, based on NOEC study on reduced leaf weight and reduced plant height (Strek & Weber,
1980 as cited in Will & Suter, 1994). Based on the data set for terrestrial plant toxicity and the
unspecified duration of the benchmark study, the terrestrial plant benchmark of 40 mg/kg was
categorized as interim.
Soil Community: Adequate data with which to derive a benchmark protective of the soil community
were not identified.
August 1995
-------
APPENDIX B
Polychlorinated Biphenyl (PBC) - Aroclor - g
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
fepimwttMiv*
" • Sptpftt
dear mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white- tailed
deer
red-tailed hawk .
American
kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil
community
Benchmark
V»(M»- mo*«-
-------
. .,•**•
APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 9
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for ecological receptor categories: trophic level 3 and 4 fish in
the limnetic and littoral ecosystems, general fish (BCF only), aquatic invertebrates, earthworms,
other soil invertebrates, terrestrial vertebrates, and plants. Each value is identified as whole-body
or lipid-based and, for the generic aquatic ecosystems, the biological uptake factors are
designated with a "d" if the value reflects dissolved water concentrations, or a "t" if the value
reflects total surface water concentrations. For organic chemicals with log K<,w values below 4,
bioconcentration factors (BCFs) in fish were always assumed to refer to dissolved water
concentrations (i.e., dissolved water concentration equals total water concentration). It should
be noted that, for the purposes of bioaccumulation modeling, a log K,jW of 6.2 was selected to
represent total PCBs based on the recommended value of 6.14 in the Great Lakes Water Quality
Initiative Technical Support Document for the Procedure to Determine Bioaccumulation Factors -
July 1994 (U.S. EPA, 1994b). The GLI technical support document calculated an arithmetic
average for the most prevalent PCB congeners. Given the level of precision required by the
models and the variability in PCB mixtures, the log Kow of 6.14 was conservatively rounded to
6.2 (i.e., two significant figures). For organic chemicals with log K^ values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using models
based on the relationship between dissolved water concentrations and concentrations in fish. The
following discussion describes the rationale for selecting the biological uptake factors and
provides the context for interpreting the biological uptake values presented in Table 4.
As stated in section 5.3.2, the BAFls for consituents of concern were generally estimated using
Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral ecosystem.
For extremely hydrophobia constituents, the Agency has stated that reliable measurements of
ambient water concentrations (especially dissolved concentrations) are not available and that
accumulation of these constituents in fish or other aquatic organisms cannot be referenced to a
water concentration as required for a BCF or BAF (U.S. EPA, 1993i). Fortunately, extremely
hydrophobic constituents can be measured in sediments and aquatic life and, because these
chemicals tend to partition to lipids and organic carbon, a biological uptake factor that reflects
the relationship between sediment concentrations and organism concentrations may be more
appropropriate. Consequently, the BSAF is the preferred metric for accumulation in the littoral
aquatic ecosystem for extremely hydrophobic chemicals (e.g., chemicals with > log Kow of - 6.5).
However, for Aroclor-1254 the predicted BAFs were used
for assessing accumulation in a littoral aquatic ecosystem because the BSAF was only available
for trophic level 4 fish and the difference between the calculated BSAF and the predicted BAF
was minimal.
For the limnetic ecosystem and the other trophic levels in the littoral ecosystem, predicted BAF^s
were used to estimate the bioaccumulation potential of PCBs in fish and aquatic invertebrates.
BSAFs were not recommended for trophic levels 3 and 2 in the littoral ecosystem and, therefore,
BCF^s from the Thomann model (1992) were used. The BAF,d for trophic level 4 fish in the
limnetic ecosystem was in good agreement with the value proposed in the GLI technical support
document (U.S. EPA, 1994b). The GLI (U.S. EPA, 1994b) value of 12,000,000, presumably
based on the analysis of measured BAFjd values, was within a factor of 2 of the predicted value
from the Thomann model (1989). The bioconcentration factor for fish was estimated from the
August 1995
-------
APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 10
Thomann models (i.e.-, log Kow ~ dissolved BCF/) and multiplied by the dissolved fraction (/j)
as defined in Equation 6-21 to determine the total bioconcentration factor (BCF/). The dissolved
bioconcentration factor (BCF/1) was convened to the BCF/ in order to estimate the acceptable
lipid tissue concentration (TC/) in fish consumed by piscivorous fish (see Equation 5-115). The
BCF/ was required in Equation 5-115 because the surface water benchmark (i.e., FCV or SCV)
represents a total water concentration (Cl). Mathematically, conversion from BCF/1 to BCF/ was
accomplished using the relationship delineated in the Interim Report on Data and Methods for
Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Wildlife (U.S. EPA, 1993i):
BCF/1 x fd = BCF/
Converting the predicted BCF/* of 1,972,422 L/kg LP to the BCF/ of 297,724 L/kg LP was in
good agreement (i.e., within a factor of ~2) with the geometric mean of 30 measured BCF/
values for different PCB congeners presented in the Derivation of Proposed Human Health
and Wildlife Bioaccumulation Factors for the Great Lakes Initiative (geometric mean =
409,414).
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of measured
values cited in Garten and Trabalka (1983). For earthworms and terrestrial invertebrates, the
bioconcentration factor was estimated as described in Section 5.3.5.2.3. Briefly, the
extrapolation method is applied to hydrophobic organic chemicals assuming that the
partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs and
BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data is compared to the BBF for TCDD and that ratio (i.e., PCBs
BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial vertebrates,
invertebrates, and earthworms, respectively. For hydrophobic organic constituents, the
bioconcentration factor for plants was estimated as described in Section 6.6.1 for above
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
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APPENDIX B
Polychlorinated Biphenyl (PBC) • Aroclor - 11
Table 4. Biological Uptake Properties
«cologlc*j
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF,8AF,or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
UpfeMNUMd or
whofo-body
lipid
lipid
lipid
lipid
lipid
lipid
whole-body
whole-body
whole- body
whole-plant
value
28,811, 266 (d)
11. 210,594 (d)
409,414 (t)
29,494.3391 (d)
30,680,700
32,153.368
3.5
2.3
18
0.0089
•ourc*
. predicted value based on
Thomann, 1989, food chain
model
predicted value based on
Thomann, 1989. food chain
model
predicted value based on
Thomann, 1989 and adjusted
to estimate total BCF
predicted value based on
Thomann, 1 989, food chain
model
predicted value based on
Thomann, 1 992, food chain
model
predicted value based on
Thomann, 1992, food chain
model
Garten and Trabalka, 1983
Cooke. 1972 as cited in
WHO, 1989
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCOD
U.S. EPA, 1992e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Polychlorinated Biphenyl (PBC) - Aroclor - 12
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Polychlorinated Biphenyls and Methylmercury, Singly and in Combination on Mink. I:
Uptake and Toxic Responses. Arch. Environ. Contain. Toxicol. 16, 441-447.
Wren, C. D., D. B. Hunter, J. F. Leatherland, and P. M. Stokes, 1987b. The Effects of
Polychlorinated Biphenyls and Methylmercury, Singly and in Combination on Mink. II:
Reproduction and Kit Development. Arch. Environ. Contain. Toxicol. 16, 449-454.
August 1995
-------
Terrestrial Toxicity - PCB - Aroclor 1254
PCB Conqener
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
CAS
Number
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
Species
European
starling
red-winged
blackbird
brown-
headed
cowbird
mallard
mouse
rat
rat
rat
rat
rat
Endpolnt
acute
acute
acute
acute
rep
dev
rep
rep
(etotox
fetotox
Description
LD50
LD50
LD50
LD50
NOAEL
LOAEL
LOAEL
LOAEL
LOAEL
NOAEL
Value
1,500
1,500
1,500
> 2,000
1.25
77
30
8
2.5
50
Units
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg -d
mg/kg-d
mg/kg -d
mg/kg-d
mg/kg-d
mo/kg-d
Exposure
Route (oral,
s.c., l.v., l.p.,
Injection)
diet
diet
diet
oral
diet
diet
gavage
oral
diet
gavaqe
-
Exposure
Duration /
Timing
4 days
6 days
7 days
single dose
108 days.
NS
1 x day for 1
month
lactating days
1,3,5,7, and 9
during
gestation
gestation days
7-15
'
Reference
Stickel et at.. 1984
as cited in Eisler,
1986
Stickel et al., 1984
as cited in Eisler,
1986
Stickel et al., 1984
as cited in Eisler,
1986
NAS, 1979 as
cited in Eisler,
1986
Welsh, 1985
Spencer, 1 982
Brenzer et al.,
1984 as cited in
ATSDR, 1993
Sageretal., 1987
Collins & Capen,
1980 as cited in
Opreskoetal ,
1993
Linderetal., 1974
as cited in
ATSDR, 1993
Comments
•
Dose of 0. 1 25, 1 .25 and 12.5
mg/kg-d were administered.
Females in the 12.5 mg/kg-d
conceived at a lower rate (55%)
than the control group
Reduced average fetal weight per
litter at birth
Increased estrus and decreased
receptivity
Doses of 0, 8 mg/kg, 32 mg/kg,
and 64 mg/kg. Decreased male
fertility and decreased number of
embryos
PCB- 1254 concentration in diet =
50 mg/kg food. Calculated daily
dose from food factor of 0.05
At 1 00 mg/kg-d dose, there was
60% decreased pup survival at
weaninq
-------
Terrestrial Toxlclty JB - Aroclor 1254
PCB Conqener
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1 254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
CAS
Number
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
Species
rat
mouse
mink
mink
mink
mink
raccoon
northern
bobwhite
mallard
ring-
necked
pheasant
Japanese
quail
Endpolnt
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
acute
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
1,010
>250
6.7
48.5
79
4,000
>50
604
2,699
1,091
2,898
Units
mg/kg-body wt.
mg/kg
mg/kg
mg/kg
ma/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mq/kq
Exposure
Route (oral,
S.C., I.V., l.p.,
Inlectlon)
oral
diet
diet
diet
diet
oral
diet
diet
diet
diet
diet
Exposure
Duration /
Timing
NS
3- 18 weeks
9 months
35-day
28-day
single dose
8 days
8 days
8 days
8 days
8 days
Reference
RTECS, 1994
Talcott and Koller,
1 983 as cited in
Eisler, 1986
Ringer, 1983 as
cited in Eisler,
1986
Aulerich et al.,
1986
Aulerich et al.,
1986
Aulerich and
Ringer, 1977;
Ringer, 1983 as
cited in Eisler,
1986
Montzetal., 1982
as cited in Eisler,
1986
Heath eta!., 1972
as cited in Eisler,
1986
Heath etal., 1972
as cited in Eisler,
1986
Heath etal., 1972
as cited in Eisler,
1986
Heath etal., 1972
as cited in Eisler,
1986
Comments
•
i
'
-------
Terrestrial Toxicity - PCB - Aroclor 1254
PCB Congener
Aroclor- 1254
Metabolized Aroclor-
1254
Metabolized total
PCB's
Aroclor- 1254
Metabolized total
PCB's
Aroclor- 1254
CAS
Number
11097-69-1
11097-69-1
11097-69-1
11097-69-1
Species
mink
mink
mink
mink
mink
rhesus
monkev
Endpolnt
rep
rep
rep, kit
growth
rep
rep. kit
growth
rep
Description
LOAEL
LOAEL
LOAEL
NOAEL
NOAEL
PEL
Value
0.375
0.096
0.072
0.15
0.032
0.28
Units
mg/kg-d
mg/kg-d
mg/kg-d
mo/kg-d
mg/kg-d
mg/kq-d
Exposure
Route (oral,
8.9., l.v., I. p.,
Injection)
diet
oral
diet
diet
diet
diet
Exposure
Duration /
Timing
12. 5 weeks
160 days
up to 290 days
4-months
up to 290 days
38 weeks
Reference
Aulerich et al.,
1985
Platonow and
Karstad, 1973 as
cited in U.S. EPA,
1993b
Homshaw et al.,
1983
Aulerich and .
Ringer, 1977
Homshaw et al..
1983
Arnold etal. 1990
Comments
Concentrations ranged from 0.1
ppm to 5.0 ppm in the diet, no live
kits wer produced at 2.5 ppm .
Aroclor 1 254 to Jersey cows and
then feeding the resulting
contaminated beef to mink over
160 days at 0.64 amd 3.57 ppm
total PCB's.
Dietary concentrations ranged
from 0.21 to 1.5 ppm, at 0.48 pprr
PCB residues, mink had inferior
reproductive performance and/or
kit survival.
Dietary doses of 1 , 5, and 15 ppm
Aroclor 1254 were administered.
Reduced reproduction at 5 ppm
and 1 5 ppm - no effect on
reproduction rate at 1 ppm dose.
Dietary concentrations ranged
from 0.21 to 1 .5 ppm, at 0.48
ppm, PCB residues mink had
inferior reproductive performance
and/or kit survival.
HII •» Ul 11 IU UUdlUU IIIHbUb
monkeys aborted within 30-60
days after becoming pregnant,
while all control monkeys had
viable offspring. Increased post-
implant bleeding also noticed in
treated monkeys.
-------
Terrestrial Toxiclty . OB - Aroclor 1254
PCB Conaener
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
^
CAS
Number
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
Species
rat
rat
rabbit
white-
footed
mice
white-
footed
mice
white-
footed
mice
mink
mink
EndDoInt
rep
rep
fetotox
rep
rep
rep
rep
dev
Description
NOAEL
NOAEL
NOAEL
PEL
LOAEL
LOAEL
PEL
PEL
Value
8
0.32
10
68
34
1.53
0.3
0.15
Units
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kq-d
Exposure
Route (oral,
S.CM l.v., l.p.,
Injection)
oral
diet
gavage
-
diet
diet
oral (diet)
diet
diet
Exposure
Duration /
Timing
lactating days
1-3,5,7, and 9
129 days
gestation days
1-28
2-3 weeks
60 days
NS
9-months
6 months
Reference
Sageretal., 1983
Under etal., 1974
as cited in -
ATSDR, 1993
Villeneuve efal.,
1972 as cited in
ATSDR, 1993
Sanders &
Kirkpatrick, 1975
as cited in
Opresko el al.,
1993
Merson &
Kirkpatrick, 1976
as cited in
Opresko et al.,
1993
Linzey, 1988 as
cited in U.S. EPA,
1993b
Aulerich and
Ringer, 1977
Wren etal.,
1987a;
Wren etal., 1987b
Comments
and 64 mg/kg. Decreased numbei
of embryos at the two higher .
doses.
1
Decreased litter size at 1 .5
mg/kg-d
71% fetal death at 12.5 mg/kg-d
PCB- 1254 concentration in diet -
400 mg/kg. Calculated daily dose
from food factor of 0. 1 7
PCB- 1254 concentration in diet =
200 mg/kg. Calculated daily dose
from food factor of 0. 1 7
Reduced reproductive organ
weights, drastically reduced
number of litters and survival
among the young of the second
generation treated group.
Dietary dose of 2 ppm Aroclor
1254 was administered, adverse
effects on reproduction include 2
of 7 females whelped and 1 live,
underweight kit was produced.
At 3 and 5 weeks, the growth rate
of kits nursed by mothers
exposed to 1 .0 ug/g Aroclor- 1254
was significantly reduced
-------
Freshwater Biological Uptake Measures - PCB - Aroclor 1254
Chemical Name
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1 254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
-
Species
blueglll
channel catfish
fish
fish
fish
fish
Daphnia magna
fish
lake frout
largemouth bass
B-factor
(BCF. BAf,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BAF
BAF
BAF
Value
26.300 -
71.400
61.900
12.023
45.709
141.254
31.200
47.000
50.119
16.218.101
5.248,075
Measured
or
Predicted
(m.p)
m .
m
P
P
P
m
P
m
P
P
units
(I/kg. NS.
other)
NS
NS
NS
NS
NS
L/kg
NS
L/kg
L/kg LP
L/kg LP
Reference
Stalling & Mayer,
1972
Mayer etal.. 1977
Kenega &
Goring. 1980 as
cited in Mackay
etal., 1992
Kenega &
Goring. 1980 as
cited in Mackay
etal., 1992
Mackay. 1982
U.S. EPA, 1992
NAS. 1979 as
cited in Eisler.l 986
Garten &
Trabalka. 1983
Thomann, 1989
momann. 1989
Comments
BCF for bluegills chronically
exposed to 2 - 10 ug/l ranged from
26.300 to 71.400 times the exposure
levels.
BCF value was measured after 77
days in the water. PCB uptake
had not reached equilibrium at the
end of exposure.
Correlated BCF value was
calculated Vieth et al.'s BCF value
of 100,000. Correlation uses BCF
and aqueous solubility.
Normalized BCF to 3% lipid
BAF value is the geometric mean
of 4 values.
-------
Terrestrial Toxiclty JB - Aroclor 1254
PCB Congener
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
CAS
Number
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
Species
cynomolgu
s monkey
mallard
chicken
chicken
chicken
chicken
pheasant
Endpolnt
fetotox
"
rep
egg
production
and fertility
chick
growth
egg
hatchability
egg
production
and
hatchability
egg
hatchability
Description
LOAEL
NOAEL
LOAEL
LOAEL
NOAEL
NOAEL
LOAEL
Value
0.1
1.45
2.44
0.98
2.44
0.96
1.8
Units
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
m
-------
Freshwater Biological Uptake Measures - PCB - Aroclor 1254
Chemical Name
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Species
fish
fish
fish
plankton to fish
B-foctor
(BCF. BAF.
BMF)
BCF
BCF
BCF
BMF
Value
33.636
2.089
13.159
12.9
Measured
or
Predicted
(m,p)
m
m
m
m
units
(I/kg. NS.
other)
NS
NS
NS
NS
Reference
Hansen et al..
1971
Snarski and
Puglisi, 1976 as
cited in Stephen.
1993
Veithetal.. 1979b
as cited in
Stephan. 1993
Evans etal.. 199)
Comments
BCF value is normalized to 1% lipid
BCF value is normalized to 1% Ijpjd
BCF value is normalized to 1% lipid
NS = not specified
-------
Freshwater Toxiclt> . CB - Aroclor 1254
'
Chemical
Name -
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
CAS
Number
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
11097-69-1
Species
Daphnia
magna
brook trout
fathead
minnow
Daphnia
magna
bluegill
rainbow
trout
aquatic
organisms
Type of
Effect
lite cycle
life cycle
life cycle
rep
acute
acute
NS
Description
cv
cv
cv
EC50
LC50
LC50 .
SCV
Value
2.1
1
2.9
1.1 -25
(11-8)
2740
0.32
0.02
Units
ug/l
ug/l
ug/l
ug/l
yg/1
ug/l
ug/l
Test Type
(static/ flow
through)
flow-through
flow-through
flow-through
NS
NS
NS
NS
Exposure
Duration /
Timing
NS
NS
NS
14-day
4-day
4-day
NS
Reference
Nebeker & Puglisi,
1974 as cited in U.S.
EPA, 1980
Maucket al., 1978 as
cited In U.S. EPA,
1980
Nebeker et al., 1974
as cited in U.S. EPA,
1980
AQUIRE, 1995
Stalling & Mayer,
1972
Birgeetal., 1978 as
cited in AQUIRE
Suter and Mabrey.
1994
Comments
.
NS = Not specified
-------
Terrestrial Biological Uptake . ^asures - PCB - Aroclor 1254
Chemical Name
Aroclor- 1254
Aroclor- 1 254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1254
Aroclor- 1 254
Aroclor- 1 254
Species
plant
sheep
\
poultry
small birds
rodents
cow
swine
B-faclor
(BCF. BAF.
BMP)
BCF
BAF
BAF
BAF
BAF
BAF
BAF
Value
0.0089
1.5'
5.9
9.5
6.2
3.4
-I.I
Measured
or
Predicted
(m,p)
P
P
P
P
P
P
P
Units
(ug/g DW
plant)/(ug/g
soil)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
Reference
U.SEPA. 1990e
Garten &
Trabalka. 1983 •
Garten &
Trabalka. 1983
Garten &
Trabalka 1983
Garten &
Trabalka 1983
Garten &
Trabalka 1983
Garten &
Trabalka, 1983
Comments
Plant uptake from soil pertains to
leafy vegetables
% lipid was not specified in study.
% lipid was not specified in sludy.
% lipid was not specified in study.
% lipid was not specified in study.
% lipid was not specified in study.
% lipid was not specified in study.
NS = not specified
-------
APPENDIX B Selenium - 1
Toxicological Profile for Selected Ecological Receptors
Selenium
Cas No.: 7782-49-2
Summary: This profile on selenium summarizes the lexicological benchmarks and biological
uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from data
presented in the technical support document for the Hazardous Waste Indentification Rule
(HW1R): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Three possible benchmark studies were identified that involved selenium toxiciry in
mammals. In one study, Schroeder and Mitchener (1971b) assessed the reproductive effects of
selenium in three generations of mice. A single dose of 3 ppm selenium was administered in
drinking water. Mice in all three generations produced fewer number of offspring and a greater
percentage of runts than the controls. The 3 ppm dose was converted to a daily dose of 0.71
mg/kg-day by using the geometric mean of the reference water intake rates (0.008 L/day) and
bodyweights (0.035 kg) for two types of typical laboratory mice (U.S.,EPA, 19881). Nobunaga
et al., (1979) exposed mice to two oral doses of selenium in drinking water for 30 days prior to
mating and for the first 18 days of gestation. No significant effects on reproduction or incidences
of fetotoxicity were evident at the lower dose of 11.4 nmol/ml (NOAEL), however, the higher
dose of 22.8 nmol/ml (LOAEL) resulted in a significant reduction in fetal growth. These effects
August 1995
-------
APPENDIX B Selenium - 2
levels correspond to daily doses of 0.9 mg/kg-day and 1.8 mg/kg-day. To arrive at these doses,
the molecular weight of sodium selenite was used to convert the nmol/ml doses to ppm doses.
The ppm dose was then converted to the daily dose by using the geometric mean of the reference
water intake rate for lab mice of 0.008 L/day (U.S. EPA, 19881) and the mice bodyweights of
0.28 kg that were given in the study. Rosenfeld and Beath (1954) examined the effects of
selenium on the reproduction of successive generations of Wistar rats. The authors administered
doses of 1.5, 2.5 and 7.5 ppm of selenium in drinking water. The 2.5 ppm dose was reported
to have reduced the number of young reared by the second generation mothers by 50%. This
reduction resulted in a LOAEL of 2.5 ppm and a NOAEL of 1.5 ppm. These effects levels
correspond to daily doses of 0.34 and 0.20 mg/kg-day based on the Wistar rat's reference
bodyweight of 0.320 kg and water consumption rate of 0.043 L/day (U.S. EPA, 1988).
The NOAEL for reproductive effects from the Rosenfeld and Beath (1954) study was chosen to
derive the lexicological benchmark because (1) chronic exposures were administered via oral
ingestion, (2) the study focused on longterm reproductive success as a critical endpoint, (3) the
study contained dose response information, and (4) the study contained the lowest toxicity value
for a critical endpoint. The Schroeder and Mitchener study (1971b) was not chosen for the
derivation of the benchmark because it did not contain sufficient dose response information. The
Nobunaga (1979) study was not chosen because it did not report the lowest toxicity value for a
critical endpoint Therefore, the NOAEL of 0.20 mg/kg-day from Rosenfeld and Beath (1954)
was chosen for the derivation of a mammalian benchmark value.
The study value from Rosenfeld and Beath (1954) was scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al. (1994):
Benchmark = NOAEL, x
*V 4
where NOAELj is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the critical endpoint
selected from the Rosenfeld and Beath (1954) study was the reproductive success of female rats,
the mean female body weight of representative species was used in the scaling algorithm to
obtain the lexicological benchmarks.
Data were available on the reproductive and developmental effects of selenium, as well as growth
or chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. There were study values in the
data set that were more than an order of magnitude below the benchmark value (Chowdhury and
Venkatakrishna-Bhatt, 1983). Based on the data set for selenium, the benchmarks developed
from the Rosenfeld and Beath (1954) study were categorized as adequate, with a "*" to indicate
that some adverse effects have been observed at the benchmark level.
August 1995
-------
APPENDIX B Selenium - 3
Birds: Only one study was identified that investigated the effects of selenium toxicity on avion
species. Mallard duck pairs were fed diets containing selenium as sodium selenite for 4 weeks
prior to egg laying at doses of 1, 5, 10, 25 and 100 ppm (Heinz et al., 1987). Although there
were no effects on the reproductive success of the adults at the 1, 5, and 10 ppm dose levels,
females fed 25 ppm took longer to begin laying eggs and intervals between eggs were longer.
This resulted in a LOAEL and a NOAEL of 25 and 10 ppm, respectively. These effects levels
correspond to daily doses of 2.5 and 1.0 mg/kg-day, converted from the ppm doses, by using the
food intake rate of 105.5 g/day and the geometric mean (1.055 kg) of the control body weights
given in the study.
The NOAEL of 1.0 mg/kg-day from the Heinz et al. (1987) study was selected to derive the
avian benchmark value for the freshwater ecosystem. This study was chosen because (1) chronic
exposures were administered via oral ingestion, (2) reproductive toxicity was one of the primary
endpoints examined, and (3) the study contained sufficient dose-response information.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian species
representative of a freshwater ecosystem, the NOAEL value of 1.0 mg/kg-day from the Heinz
et al. (1987) study was scaled using the cross-species scaling method of Opresko et al. (1994).
Since the reproductive endpoint examined in the Heinz et al. (1987) study entailed dosing male
and female mallards, both male and female body weights for each representative species were
used in the scaling algorithm to obtain the lexicological benchmarks.
Data were available on reproductive and developmental effects of selenium as well as on growth
and survival. In addition, the data set contained studies that were conducted over chronic and
subchronic durations as well as during a sensitive life stage. There were no other values in the
data set that were an order of magnitude or more below the benchmark value. Based on the
avian data set for selenium, the benchmarks developed from the NOAEL in the Heinz et al.
(1987) study were categorized as adequate.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 5.0E-03 mg/L for selenium
was selected as the benchmark protective of fish and aquatic invertebrates (U.S. EPA, 1987).
Since the FCV was derived in the AWQC document, the benchmark was categorized as
adequate.
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g. duckweed) or (2) an effective concentration (ECXX) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastrum capricornutum). For selenium the
benchmark value presented by Suter and Mabrey (1994) was l.OE+02 ug/L based on the growth
inhibition of Scenedesmus oblicuus. As described in Section 4.3.6, all benchmarks for aquatic
plants were designated as interim.
August 1995
-------
APPENDIX B
Selenium - 4
Benthic community: The selenium benchmark protective of benthic organisms is pending a U.S.
EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
ReprennUitfw
ft***
mink
river oiler
bald eagle
osprey
great blue heron
mallard
lesser scaup
•potted sandpiper
herring gut
kingfisher
S*K*«Ult
vnfr***fl*8y
***
0.17(a')
0.09 (a*)
0.73 (a)
0.90 (a)
6.82 (a)
0.98 (a)
1.09 (a)
2.'23 (a)
0.99 (a)
1.64 (a)
SON*
$p*ci«»
rat
rat
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
mallard duck
EH**
rep
rep
rep
rep
rep
rep
rep
rep
rap
rep
wa*Mwf
0.20
0.20
1.0
1.0
1.0
1.0
1.0
1.0
1.0
1.0
QeeqtoUuu
•c ~~
NQAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
tr
•
-
• -
.
-
.
-
-
ftAtfeuftt ftoitf i*tft
Rosenfeldand
Beaih. 1954
Rosen told and
Beatfi, 1954
Heinz et al., 1987
Heinz et al., 1987
Hevuetal., 1987
Heinz etal.. 1987
Heinz etal., 1987
Heinz et al., 1987
Heinz etal., 1987
Heinz et al., 1987
•Benchmark Category, a * adequate, p - provisional, i = interim; a "' indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Selenium - 5
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
•XT"
fish and aquatic
invertebrates
aquatic plants
benlhic community
Value*
5.0E-03 (a)
1.0E+02ug/L(i)
under review
»
aquatic
organisms
aquatic
plants
-
—
FCV
CV
-
*»,»«.
AWQC
Suter and Mabrey,
1994
-
OA*vfcm*flr f^atotfuw A — *r4omiatA n - nmuiunrukl i — inform* A •*' inriirfltoc tHot tHa hfinrfima^r
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Selenium-6
II. lexicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C-J for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to selenium.
Because of the lack of additional mammalian toxicity studies, the same surrogate species
study (Rosenfeld and Beath, 1954) was used to derive the selenium lexicological benchmark
for mammalian species representing the terrestrial ecosystem. The study NOAEL of 1.5 ppm
(0.20 mg/kg-day) was scaled for species in the terrestrial ecosystem using a cross-species
scaling algorithm developed by Opresko et al. (1994). Since the Rosenfeld and Beath (1954)
study documented reproductive effects from selenium exposure to female rats, the female
body weight of each representative species was used in the scaling algorithm to obtain the
lexicological benchmarks.
Based on the data sei for selenium, the benchmarks developed from the Rosenfeld and Beath
(1954) study were categorized as adequate, with a "*" lo indicate lhai some adverse effecis
have been observed al ihe benchmark level.
Birds: As in the freshwater ecosystem, the sludy by Heinz el al. (1987) was used lo calculate
ihe benchmarks for birds in ihe generic lerresirial ecosystem. The study NOAEL of 10 ppm
(1.0 mg/kg-day) was scaled for the representative species by using the cross-species scaling
algorithm developed by Opresko et al. (1994). Since the reproductive endpoini examined in
the Heinz et al. (1987) study entailed dosing male and female mallards, both the male and
female body weights for each represenlative species were used in the scaling algorithm lo
oblain the lexicological benchmarks. Based on the avion data sei for selenium, the
benchmarks developed from ihe Heinz el al. (1987) sludy were categorized as adequate.
Plants: Adverse effecis levels for terrestrial plants were identified for endpoints ranging from
perceni yield lo rool length. As presented in Will and Suter (1994), phyioloxicity benchmarks
were selected by rank ordering the LOEC values and then approximating ihe 10 percentile.
If ihere were 10 or fewer values, ihe 10th percentile LOEC was used. Such LOECs applied lo
reductions in planl growih, yield reductions, or oiher effecis reasonably assumed lo impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation. The
selected benchmark for phyioioxic effecis of selenium in soils is 1.0 mg/kg (Will and Suler,
1994). Since the study value selected is ihe 10th percentile of more lhan 10 LOEC values, ihe
terrestrial benchmark for selenium is categorized as provisional.
August 1995
-------
APPENDIX B Selenium-7
Soil Community: Adequate data with which to derive a benchmark protective,of the soil
community were not identified.
August 1995
-------
APPENDIX B
Selenium - 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
XT"*
deer mouse
short-tailed .
shrew
meadow vote
Eastern
cottontail
red fox
raccoon
white- tailed dear
red- tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
0.40 (a*)
0.42 (a*)
0.34 (a*)
0.1 4 (a*)
0.11 (a*)
OIIO (a*)
0.05 (a*)
0.98 (a)
173 (a)
1.57 (a)
1.90 (a)
1.58 (a)
1.0 (p)
mg/Kg
ID
Sledy
rat
rat
rat
rat
rat
rat
rat
mallard
duck
malard
duck
mallard
duck
mallard
duck
malard
duck
terrestrial
plants
EHeot
rep
rep
rep
rep
rep
rep
rep
dev
dev
dev
dev
dev
growth/
yield
•
««*(*>
0.20
0.20
0.20
0.20
0.20
0.20
0.20
1.0
1.0
1.0
1.0
1.0
1.0 mgAg
-
*-*- .
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
10th percentile
- LOEC
-
w
••
•
•
•
•
-
-
•
•
•
•
•
-
Rosen feld and
Bealh, 1954.
Rosenfeldand
Bealh. 1954
Rosen feld and
Beath, 1954
Rosen feld and
Bealh, 1954
Rosen feld and
Bealh. 1954
RosenMdand
Bealh, 1954
RosenMdand
Bealh, 1954
Heinz et al., 1987
Heinz et al.. 1987
Heinz et al., 1987
Heinz et al., 1987
Heinz et al., 1987
Will and Suter,
1994
-
'Benchmark Category, a - adequate, p = provisional, i = interim; a "" indicates (hat the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.
ID = inouffoent data
August 1995
-------
APPENDIX B Selenium. 9
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: general fish
(BCF only), aquatic invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates,
and plants. For metals, BCFs are whole-body bioconcentration factors and refer to total
surface water concentrations (versus freely dissolved concentrations). Consequently all
calculations of acceptable tissue concentrations (TC) represent whole-body concentrations.
The following discussion describes the rationale for selecting the biological uptake factors
and provides the context for interpreting the biological uptake values.
The whole-body BCF for selenium was the geometric mean of several measured values,
from several sources (e.g. Besser et al., 1993). The geometric mean of 88 was calculated
from 6 sources which presented values ranging from 2-918. BCF values for muscle were not
included because ecological receptors are likely to eat the whole fish, or in least, will not
necessarily distinguish between the fillet and other parts of the fish. Data on bioconcentration
in aquatic invertebrates are under review. Appropriate studies on bioconcentration/
bioaccumulation were not identified for terrestrial vertebrates and invertebrates (including
earthworms). The whole-plant BCF value was determined to be 6.0E-03 (U.S. EPA, 1992e).
For metals, empirical data were used to derive the BCF for aboveground forage grasses and
leafy vegetables. In particular the uptake-response slope for forage grasses was used as the
BCF for plants in the terrestrial ecosystem since most of the representative plant-eating
species feed on wild grasses.
August 1995
-------
APPENDIX B
Selenium - 10
Table 4. Biological Uptake Properties
•OOipgfeaf
receptor
fish
littoral trophic
level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BAF,or
BSAF
BCF
•
-
BCF
ttpHMWMdQf
whole-tody
whole
•
•'
•
whole-plant
VeilM
88 (t)
ID
ID
ID
ID
6.0E-03
•awto
geometric mean of several
measured values for whole
body BCFs as cited in the
master table (e.g., Better etal...
1993)
•
'
•
U.S. EPA. 1992e
d » rotor* to dissolved surface water concentration
t - rotors to total surface water concentration
10 - insufficient data
August 1995
-------
APPENDIX B Selenium - 11
References
AQUIRE (AQf/atic Toxicity /nformation /JEtrieval Database). 1994, Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Barrows, M. E., S. R. Petrocelli, and K. J. Macek. 1980. Bioconcentration and elimination
of selected water pollutants by bluegill sunfish (Lepomis macrochirus). In R. Haque (ed.).
Dynamics, Exposure and Hazard Assessment of Toxic Chemicals. Ann Arbor Science
Pub. Inc., Ann Arbor, MI. pp. 379-392.
Besser, J. M., T. J. Canfield, and T. W. LaPoint. 1993. Bioaccumulation of organic and
inorganic selenium in a laboratory food chain. Environmental Toxicology and Chemistry
12:57-72. .
Chowdhury, A. R., and H. Venkatakrishna-Bhatt. 1983. Effect of selenium dioxide on the
testes of rat Indian J Physiol Pharmacol. 27:237-240.
Coyle, J. J., D. R. Buckler, C. G. Fairchild, and T. W. May. 1993. Effect of dietary
selenium on the reproductive success of bluegills (Lepomis macrochirus). Environmental
Toxicology and Chemistry 12:551-565.
Eisler, R. 1985. Selenium Hazards to Fish, Wildlife and Invertebrates: a Synoptic Review.
U.S. Fish and Wildlife Service. Biological Report 85(1.5).
Perm, V. H., D. P. Hanlon, C. C Willhite, W. N. Choy, S. A. Book. 1990. Embryotoxicity
and dose-response relationships of selenium on hamsters. Reproductive Toxicity 4:183-
190.
57 FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg3/4/day.
Hamilton, S. J., and K. J. Buhl. 1990. Acute toxicity of boron, molybdenum, and selenium
to fry of chinook salmon and coho salmon. Archives of Environmental Contamination
and Toxicology 19(3):366-373.
Harr, J. R., and O: H. Muth. 1972. Selenium poisoning in domestic animals and its
relationship to man. Clin Toxicol 5:175-186. As cited in Toxicological Profile for
Selenium^ Agency for Toxic Substances and Disease Registry, U.S. Public Health Service,
1989.
August 1995
-------
APPENDIX B Selenium-12
Heinz, G. H., D. J. Hoffman, A. J. Krynitsky, and M. G. Weller. 1987. Reproduction in
mallards fed selenium. Environmental Toxicology and Chemistry 6:423-433.
Hodsen, P. V., D. J. Spry, B. R. Blunt. 1980. Effects on rainbow trout (Salmo Gairdneri) of
a chronic exposure to waterborn selenium. Can. J. Fish. Aquat. Sci. 37:233-240.
Lemly, D. A. 1985. Toxicology of selenium in a freshwater reservoir: implications for /
environmental hazard evaluation and safety. Ecotoxicology and Environmental Safety
10:314-338.
National Institute for Occupational Safety and Health. RTECS (Registry of Toxic Effects of
Chemical Substances) Database. March 1994.
Nobunaga, T., H. Satoh, and T. Suzuki. 1979. Effects of sodium selenite on methylmercury
embryotoxicity and teratogenicity in mice, toxicol Appl Pharmacol 47:79-88.
NTP. 1980c. Bioassay of Selenium Sulfide (Gavage) for Possible Carcinogenicity.
Bethesda, MD: National Toxicology Program, National Cancer Institute, National
Institutes of Health, NCI Technical Report Series No. 194, NTP No. 80-17. As cited in
Toxicological Profile for Selenium,-Agency for Toxic Substances and Disease Registry,
U.S. Public Health Service, 1989.
Ohlendorf, H. M., A.W. Kilness, J. L. Simmons, R. K. Stroud. 1988. Selenium toxicosis in
wild aquatic birds. Journal of Toxicology and Environmental Health. 24:67-92.
Ohlendorf, H. M., J. P. Skorupa. 1989. Selenium in relation to wildlife and agricultural
drainage water. In: Proceedings of the Fourth International Symposium on the Uses of
Selenium and Tellurium.
Ohlendorf, H. M., R. L. Hothem, C. M. Bunck, K. C. Marois. 1990. Bioaccumulation of
selenium in birds at Kesterson Reservoir, California. Arch. Environ. Contam. Toxicol.
19:495-507.
Opresko D.M., B.E. Sample, and G.W. Suter II. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1.
Palmer, I. S., and O. E. Olson. 1974. Relative toxicities of selenite and selenate in the
drinking water of rats. J Nutr 104:306-314. As cited in Toxicological Profile for
Selenium, Agency for Toxic Substances and Disease Registry, U.S. Public Health Service,
1989.
Rosenfeld, I. and O.A. Beath. 1954. Effect of selenium on reproduction in rats. Proc. Soc.
Exp. Biol. Med. 87:295-297. As cited in Integrated Risk Information System (IRIS) for
Selenium and Compounds, August 23, 1993.
August 1995
-------
APPENDIX B Selenium-13
Schroeder, H. A., and Mitchener, M. 197la. Selenium and tellurium in rats: effects on
growth, survival, and tumors. J.Nutr 101:1531-1540.
Schroeder, H. A., and M. Mitchener. 197 Ib. Toxic effects of trace elements on reproduction
of mice and rats. Arch Environmental Health 23:102-106.
j
Suter n, G.W., and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-96/R1.
Suter H, G. W., M. A. Futrell, and G. A. Kerchner. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects of Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
Tarantal, A.F., C.C. Willhite, B.L. Lasley et al., 1991. Developmental toxicity of L-
selenomethionine in Macaco fascicularis. Fund. Appl. Toxicol. 16:147-160. As cited in
Integrated Risk Information System (IRIS) for Selenium and Compounds, August 23,
1993.
U.S. EPA (Environmental Protection Agency). 1987. Ambient Water Quality Criteria for
Selenium. U.S. Environmental Protection Agency, Washington, DC. Publication No.
EPA-440/5-87-006. As cited in Suter H, G. W., M. A. Futrell, and G. A. Kerchner.
1992. Toxicological Benchmarks for Screening of Potential Contaminants of Concern for
Effects of Aquatic Biota on the Oak Ridge Reservation, Oak Ridge, Tennessee. DE93-
000719. Office of Environmental Restoration and Waste Management, U.S. Department
of Energy, Washington, DC.
U.S. EPA (Environmental Protection Agency). 19881. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. EPA P338-179874. U.S.
EPA, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). /1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA. (Environmental Protection Agency). 1993. Integrated Risk Information System.
July.
U.S. Public Health Service. 1989. Toxicological Profile for Selenium. Agency for Toxic
Substances and Disease Registry.
Will, M.E., and G.W. Suter, II. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
August 1995
-------
Terrestrial To/..~.iy - Selenium
Cas No. 7782-49-2
Chemical Name
selenium
selenium
selenium
selenium
selenium
selenium ~
selenium
selenium (selenale)
selenium (selenale)
-
Species
C ..- .
at
cattle,
sheep,
lorses
macaques
monkeys^
rats
rats
rats
rats
rat
rat
Type of
Effect
acute
ep
dev
rep
r
rep
dev
dev
mortality
Description
LD50
LOAEL
NOAEL_
NOAEJ.
LOAEL
NOAEL
LOAEL
LOAEL
mortality JNOAEL
Valu§
6700
Q-L15
0.3
0.2
0.34
0.01
0.03
1.1
0.53
Units
mg/kg-body
wt.
mg/kg/day
PPJP
mg/kg;day
mg/kg-day
mg/kg/day
.
mg/kg/day
mg/kg/day
1 mg/kg/day
Exposure
Route (oral,
s.c., i.v., i.p.,
Injection)
oral
oral
nasogastric
intubation
NS
NS
i.p.
"P
oral
oral
Exposure
Duration
/Timing
NS
NS
through
gestation days
20-50__
2 generations
2 generations
90 days
90 days
4-6 weeks
4-6 weeks
Reference
RTECS, 1994
Harr and Muth, 1972 as cited
n ASTDR. I9!9
Tarantal et al., 1991 as cited
in IRIS. 1993
Rosenfeld and Beath, 1954
Rosenfeld and Beath, 1954
Chowdhury and
Venkatakrishna-Bhatt, 1983
Chowdhury and
Venkatakrishna-Bhatt, 1983
Palmer and Olson, 1974
Palmer and Olson, 1974
Comments
These levels in the diet caused
decreased conception rates and
increased fetal resorption rates.
There were no significant maternal
or fetal developmental effects or
teratogenesis found at this dose
level
No effect was observed on
reproduction, the number of young
reared or on the reproduction of
two successive generations of
dams and sires in groups receiving
1 .5 ppm.
At 2.5 ppm, there was a reduction
in the number of young reared.
No developmental effects were
observed at this dose level. (0.003
mg/day)
Partial degeneration of the
seminiferous tubular diameter and
normal Leydig cells was observed
at this dose level. (0.006 mg/day)
1 out ot 6 males died. (6 ppm)
No mortality occurred at this dose
level.
-------
Terrestrial Toxicity - Selenium
Cas No. 7782-49-2
Chemical Name
selenium (selenite)
selenium (selenite)
selenium (selenite)
selenium (selenate)
selenium (selenite)
selenium (selenite)
selenium (selenium
sul(ide)
selenium (selenium
sullide)
selenium (selenium
sulfide)
selenium (selenium
sulfide)
Species
rat
rat
rat
mouse
mouse
mouse
rat
rat
rat
mouse
Type of
Effect
mortality
mortality
mortality
dev, rep
dev
dev
mortality
mortality
mortality
mortality
Description
NOAEL
LOAEL
PEL
PEL
LOAEL
NOAEL
LOAEL
LOAEL
NOAEL
LOAEL
Value
0.53
1.1
0.25
0.71
2.25
4.5
112
56
31.6
805
Units
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
mg/kg/day
Exposure
Route (oral,
S.C., I.V., i.p.,
injection)
oral
oral
oral
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration .
/Timing
4-6 weeks
4-6 weeks
58 days
3 generations
30 days before
mating and 18
days during
pregnancy'
30 days before
mating and 1 8
days during
pregnancy
17 days
1 7 days
1 3 weeks
1 7 days
Reference
Palmer an'd Olson, 1974
Palmer and Olson, 1974
Schroeder and Mitchener,
1971a
Schroeder and Mitchener,
1971 b
Nobunagaetal., 1979
Nobunaga et al., 1979
NTP, 1980 as cited in
ASTDR.J989
NTP, 1980 as cited in
ASTDR, 1989
NTP, 1980 as cited in
ASTDR, 1989
NTP. 1980 as cited in
ASTDR, 1989
Comments
Mo mortality occurred at this dose
level.
4 out of 6 males died. (6 ppm)
50% mortality to male rats in 58
days was observed at this dose
level
50% reduction in the number of
offspring was observed at this
dose level.
Reduced fetal growth was
observed at this dose level.
No teratogenic effects were seen
at this dose level.
50% of the males died.
50% of the females died.
50% of the males died.
-------
Terrestrial Toxu-.iy - Selenium
Cas No. 7782-49-2
Chemical Name
selenium (selenium
sulfide)
selenium
selenium
selenium
selenium
(selenomelhionine)
selenium (sodium
selenile)
selenium (sodium
selenile)
NS = Not specified
Species
mouse
chickens
Japanese
quail
ducks
mallard
ducks
mallard
ducks .
mallard
ducks
Type of
Effect
mortality
rep
rep
rep.dev
rep
dev
dev
Description
LOAEL
LOAEL
LOAEL
LOAEL
FEL
NOAEL
LOAEL
Value
•
>464
7
6
300
1
0.5
1
Units
mg/kg/day
PPm_.
PPm
ppb
mg/kg-day
mg/kg-day
mg/kg-day_
Exposure
Route (oral,
s.c., i.v., i p.,
injection)
oral
oral
oral
oral (drinking
water)
oral
oral
oral
Exposure
Duration
/Timing
17 days
NS
_NS.
NS
4 weeks prior to
egg laying
4 weeks prior to
egg laying
4 weeks prior to
egg laying
Reference
NTP, 1980 as cited in
ASTDR.1989
Ort and Latshaw, 1 978 as
cited jr^FWS, 1985
EI-Bergearmi et al., 1977 as
cited in FWS, 1985
Ohlendorf et al., 1986 as cited
in FWS, 1985
Heinz etal., 1987
Heinz et al.,1987
HeinzetaL_.19B7 .
Comments
50% of the females died.
Reduced hatching of eggs was
recorded at this dose level.
Reduced hatching of eggs was
recorded at this dose level.
Resulted in poor reproduction and
developmental abnormalities in '
aquatic nesting birds, due to
interference with their reproductive
processes.
Very low hatching success was
observed at this single dose level.
No embryotoxic effects were
observed at this dose level.
Embryotoxic effects such as
stunted growth, swollen necks,
edema, and fewer than normal
feathers were observed at this
dose level.
-------
Freshwater Toxicity - Selenium
CasNo. 7782-49-2
Chemical Name
selenium
selenium
selenium
selenium
selenium
selenium
selenium
selenium
NS = Not specified
Species
alhead minnow
bluegill
aquatic
organisms
aquatic
organisms
lish
daphnid
fish
daphnid
Type of Effect
acute
dev, rep
chron
chron
chron
chron
acute
acute
Description
LC50
LOEC
AWQC
AWQC
CV
CV
EC20
EC20
Value
1000
33.3
5
35
88.32
91.65_
40
25
Units
ug/L
ug/g-
body
wt.
HB/k..
ug/L
ug/L
49/k
ug/L
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
NS
NS
NS
Exposure
Duration
timing
4 days
140 days
NS
NS
NS
NS •__
NS
NS
Reference
Halter et al . 1980 as cited in
AQUIRE , 1994
Coyleetal . 1992
53JR 177 (Jan.J5.J988)
U.S. EPA, 1987 as cited in
Suteretal., 1992
Suteretal., 1992
Suteretal.. 1992
Suteretal , 1992
Suteretal., 1992
Comments
In addition to dietary exposure,
adults were exposed to
background concentrations of 1 0
ug Se/L. Survival of try was
severely reduced.
This AWQC value is reported in
IRIS, 1993 and the Federal
Register.
Unable to explain this value;
does not seem to be based on a
residue value.
•--
-- .
Selenii - - Page 9
-------
Freshwater Biological Uptake Measures - Selenium
Cas No. 7782-49-2
Chemical Name
selenium
selenium
selenium
selenium
selenium
selenium (Se-methionine)
selenium (Se-methionine)
selenium (Se-methionine)
selenium (Se-methionine)
Species
fathead
minnow
fathead
minnow
rainbow trout
fathead
minnow
bluegill
daphnjd
daphnid
daphnid
daphnid
B-factor
(BCF. BAF.
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BAF
Value
527.00
2.083.00
2.00
12.00
2(H)0
382,000.00
229.000.00^
30.300.00
149,000.00
Measured
or
Predicted
(m,P)
m
m
-
m
m
m
m
m
m.
m
Units
ug
ug
NS
NS
NS
NS
NS
NS
Ukg
Reference
Lemly, 1985
Lemly, 1985
Barrows et al., 1980 as
cited jn U.S. EPA, 1993b
Barrows et al., 1980 as
cited in U.S. EPA. 1993b
Barrows et al., 1980 as
cited in OS. EPA, 1993b
Besser etal., 1993
Besser etal., 1993 .
Besser et al.. 1993
Besser et al., 1993
Comments
At 10ugSe/L; white
skeletal muscle.
At 10 ug Se/L; viscera.
Muscle BCF.
Muscle BCF.
Whole body BCF.
At .1 microgram Se/L.
At 1 .0 microgram Se/L.
At 10 rhicrograms Se/L.
96-hour BAF; Aqueous
exposure to .1 ug Se/L.
-------
Freshwater Biological Up.^Ke Measures - Selenium
Cas No. 7782-49-2
Chemical Name
selenium
selenium
selenium
selenium
selenium
selenium
selenium
selenium
selenium
Species
benthic insects
molluscs .
crustaceans
annelids
periphyton
largemouth
bass
largemouth
bass
carp
carp
B-factor
(BCF, BAF,
BMF)
BCF
BCF
BCF
BCF
BCF
BCF
BCF.
BCF
BCF
Value
1,395.00
817.00
790.00
1,054.00
519.00
2,019.00
3,975.00
918.00
2,891.00.
Measured
or
Predicted
(m,p)
m
m
m
m
m
m
m
m
m
Units
!£...„
L/9
L/g .
!^9
Ug
Ug
us.....
.m.
LAj
Reference
Lemly,1985^
Lemly.1985
Lemly.1985_
Lemly, 1985
Lemly, 1985
Lemly, 1985
Lemly. 1985
Lemly, 1985
Lemly, 1985
Comments
At 10 ug Se/L.
At 10 ug Se/L.
At lOug Se/L.
At 10ugSe/L.
At 1 Dug Se/L.
At 10 ug Se/L; white
skeletal muscle.
At 10 ug Se/L; viscera.
At 1 0 ug Se/L; white
skeletal muscle.
At 10 ug Se/L; viscera.
-------
Freshwater Biological UpiaKe Measures - Selenium
Cas No. 7782-49-2
Chemical Name
selenium (Se-methionine)
selenium (Se-methionine) '
selenium (Se-meJhiqnjne)
selenium (Se-methionine)
selenium (Se-methionine)
selenium (Se-methionine)
selenium (Se-methionine)
selenium (selenate/selenite)
selenium (selenite)
Species
daphnid
daphnjd
daphnid
daphnid
daphnid
blueglll
bluegill
bluegill
rainbow trout
B-factor
(BCF, BAF,
BMP)
BAF
BAF
BAF _
BAF
BAF
BCF
BCF
BCF
BCF
Value
102,000.00
14,800.00
3300
1600^
1210
8,000.00 J
5,000.00
56.00
8.30
Measured
or
Predicted
(m,p)
m
m
m
01
m
P
P.- ...
c
NS
Units
L/kg
L/kg
Ukg
MM. „.
M!<9.
NS
NS
NS
NS
Reference
Besseretal., 1993
Besseretal., 1993
Besseretal., 1993
BesseretaJ., 1993
Besseretal. ,-1993
Besseretal., 1993
Besseretal., 1993
Bessere^aUUigs
Hodson et al., .1980 as
cited in Besser et al.,
1993
Comments
96-hour BAF; Aqueous
exposure to 1 ug Se/L.
96-hour BAF; Aqueous
exposure to 10 ug Se/L.
14-day BAF; Aqueous
exposure to .1 ug Se/L.
14-day BAF; Exposure
from algae dosed with 1
ug Se/L.
14-day BAF; Exposure
from algae dosed with 10
ug Se/L.
At 1 .0 microgram Se/L.
At 10 micrograms Se/L.
At 10 micrograms Se/L.
-------
Freshwater Biological Uptake Measures - Selenium
Cas No. 7782-49-2
Chemical Name
selenium (selenate/selenite)
selenium (Se-methionine)
selenium (Se-methionine)
selenium (selenate)
selenium (selenate)
selenium (selenate)
selenium (selenate)
selenium (selenate)
Species
fathead
minnow
bluegill
bluegill
daphnid
daphnid
daphnid
daphnid
daphnid
B-factor
(BCF, BAF,
BMP)
BCF
BCF
BCF
BCF
BCF
BCF
BAF
BAF
Value
12-29
4,900.00
4,500.00
293.00
168.00
65.10
270.00
151.00
Measured
or
Predicted
(m,p)
NS
m
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS _
NS
LM
L/kg
Reference
Bertram & Brooks, 1983
as cited in Besser et al.,
1993; Adams, 1976 as
cited in Besser et al.,
1993
Besser et al., 1993
Besser el al, 1993
Besser etal., 1993
Besser et al., 1993
Besser etal., 1993
Besser etal., 1993
Besser etal., 1993
Comments
at 1.0 microgram Se/L
at 10 micrograms Se/L
at 10 micrograms Se/L
at 100 micrograms Se/L
at 1000 micrograms Se/L
96-hour BAF; Aqueous
exposure to 10 ug Se/L
96-hour BAF; Aqueous
exposure to 1 00 ug Se/L
i
-------
Freshwater Biological Up^e Measures - Selenium
Cas No. 7782-49-2
Chemical Name
selenium (selenate)
selenium (selenate)
selenium (selenate)
selenium (selenate)
selenium (selenate/selenite)
selenium (selenate/selenite)
selenium (selenite)
selenium (selenite)
selenium (selenite)
Species
daphnid
daphnid
daphnid
daphnid
bluegill
bluegill
daphnid
daphnid
daphnid
B-factor
(BCF, BAF,
BMP)
BAF
BAF
BAF
BAF
BCF
BCF
BCF
BCF
BCF
Value
65.20
110.00
23.00
23.10
56.00
20.00
3.650.00
570.00
221.00
Measured
or
Predicted
(m,p)
m
m _
m
m
P
P_
m
m
m
Units
U*9
U*9..___
LAg
L/kg
NS
NS
NS _
NS
NS
Reference
•
Besseretal , 1993
Besseretal.. 1993
Besseretal., 1993
Besseretal., 1993
Besseretal., 1993
Besseretal.. 1993
Besseretal.. 1993
Besser et al., J993
Besseretal., 1993
Comments
96-hour BAF; Aqueous
exposure to 1 ,000 ug
Se/L
14-day BAF; Exposure
from algae dosed with 10
ug Se/L
14-day BAF; Exposure
from algae dosed with
100ug Se/L
14-day BAF; Exposure
Irom algae dosed with
1 ,000 ug Se/L
at 10 micrograms Se/L;
BCF estimates for the
two Se species were
averaged
at 100 micrograms Se/L;
BCF estimates for the
two Se species were
averaged
at 1 .0 micrograms Se/L
at 1 0 micrograms Se/L
at 100 micrograms Se/L
-------
Freshwater Biological Uptake Measures - Selenium
Cas No. 7782-49-2
Chemical Name
selenium (selenite)
selenium (selenite)
selenium (selenite)
selenium (selenite)
Selenium (Selenite)
Selenium (Selenite)
NS= Not specified
Species
daphnid
daphnid
daphnid
daphnid
daphnid
daphnid
B-factor
(BCF, BAF,
BMP)
BAF
BAF
BAF
BAF
BAF
BAF
Value
3,200.00
59000^
218.00
1,200.00
160.00
90.00
Measured
or
Predicted
(m.p)
m
m
m
m
m
m
Units
L/kg
LAg
LAg __
Lfcg
Ukg
Lfcg .:.
Reference
Besser etal., 1993
Besseretal., 1993
Besser etal., 1993
Besseretal., 1993
Besser et al , 1993
Besseretal '., 1993
Comments
96-hour BAF; Aqueous
exposure to 1 ug Se/L
96-hour BAF; Aqueous
exposure to 10 ug Se/L
96-hour BAF; Aqueous
exposure to 100 ug Se/L
14-day BAF; Exposure
from algae dosed with 1
ugSe/L
14-day BAF; Exposure
from algae dosed with 10
ug Se/L
14-day BAF; Exposure
from algae dosed with
100ug Se/L
-------
Terrestrial Biological Up,_..d Measures
Cas No. 7782-49-2
Selenium
Chemical Name
selenium
Species
plant
B-factor
(BCF. BAF.
BMP)
BCF
Value
6.2
Measured
or
Predicted
(m,p)
units
(ug/g DW
planl)/(ug/g
p isoii)
Reference
U.S. EPA, 1990 as cited in
RTI, 1994
Comments
Selenium - Page 1 /
-------
APPENDIX B Silver - 1
Toxicological Profile for Selected Ecological Receptors
Silver
Cas No.: 7440-22-04
Summary: This profile on silver summarizes the toxicological benchmarks and biological uptake
measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire toxicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
, data presented in the technical support document for the Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind toxicological benchmarks used to derive protective
media concentrations (CL^,) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were identified which studied the effects
of silver toxicity on reproductive or developmental endpoints in mammalian species.
Birds: No suitable subchronic or chronic studies were identified which studied the effects of
silver toxicity in avian species.
Fish and aquatic invertebrates: No AWQC or Final Chronic Value (FCV) was available for
silver. Therefore, a Secondary Chronic Value (SCV) of 3.6 E-04 mg/1 as reported by Suter and
Mabrey (1994) was utilized. Because the benchmark selected is based on a SCV, rather than an
FCV, it was categorized as interim.
August 1995
-------
APPENDIX B
Silver - 2
Aquatic Plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or 2) an effective concentration (ECXX) for a species of freshwater algae,
frequently a species of green algae (e.g., Selenastrum capricornutwri). The aquatic plant
benchmark for silver is 30 mg/1 based on growth inhibition of Chlorella vulgaris (Suter and
Mabrey, 1994). As described in Section 4.3.6, all benchmarks for aquatic plants were designated
as interim.
Benthic community: The silver benchmark protective of benthic organisms is pending a U.S.
review of the acid volatile sulfide (AVS) methodology proposed for metals.
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
R«t»r**«al*tiv»
$f»cte*
mink
river otter
bald 0a0l«
osprey
great blu« heron
mallard
lesser scaup
spotted sandpiper
herring gui
kingfisher
ftancfurarfc
VautfjttgAQ-
4*y
ID
ID
ID
ID
ID
ID
ID
ID
-ID '
ID
SUM**
Spteftt
-
-
-
-
-
-
•
-
•
•
€««cl
•
-
-
-
•
•
-
Study Value
mfl/ko-day
-
-
•
Dnertptfan \
-
-
-
-
-
• .
-
1 , **
•
-
-
-
-
-
•
Qrffl&t«t$ouro«
-
-
-
-
•
-
•
-
-
-
'Benchmark Category, a = adequate, p = provisional, i = interim; ID = insufficient data; a (*) indicates (hat the benchmark
value was an "order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B
Silver - 3
Table 2. lexicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
Rapr***n4»tiv*
SpoctM
fish and aquatic
invertebrates
aquatic plants
benthic community
8«ncfun«fit
V«ue*
mtft
3.6 E-04 (i)
0.030 (i)
under review
Study
$P*C)4*
aquatic
organisms
aquatic
plants
-
Original
Valu*
mg/L
3.6 E-04
0.030
-
Description
scv
cv
-
OrJgM Soarw
Sutar & Mabrey,
1994
Sutar & Mabrey,
1994
-
IL
'Benchmark Category, a - adequate, p = provisional, i » interim; ID = insufficient data; a (*) indicate* tttai the benchmark
value wa* an order of magnitude or more above the NEL or LEL for other adverse effects.
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem '
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C_) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were identified which studied the toxicity effects of silver on
mammalian reproductive or developmental endpoints.
Birds: As mentioned in the freshwater ecosystem discussion, no suitable studies were
identified which investigated the effects of silver toxicity in avian species.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the Lowest Observable Effects Concentration (LOEC) values
and then approximating the 10th percentile. If there were 10 or fewer values for a chemical,
the lowest LOEC was used. If there were more than 10 values, the 10th percentile LOEC was
used. Such LOECs applied to reductions in plant growth, yield reductions, or other effects
reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. The benchmark for terrestrial plants was based on a LOEC of 2
nig/kg resulting in unspecified toxic effects on plants (Kabata-Pendias and Pendias, 1984 as
cited by Will and Suter (1994)). As less than 10 studies were presented in Will and Suter
(1994), the terrestrial plant benchmark of 2 mg/kg was categorized as interim.
August 1995
-------
APPENDIX B Silver - 4
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not available.
August 1995
-------
APPENDIX B
Silver - 5
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
JfoprMwntatfv*
$P«$fa*
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white- tailed deer
red-tailed hawk
American kestrel
Northern
bobwhitB
American robin
American
woodcock
plants
soil community
Benchmark
VatM*"
mgfltpity
ID
ID
10
ID
ID
ID
ID
ID
ID
ID
ID
ID
2 mg/kg (i)
ID
SUMly
8p*c$«»
-
-
-
:-
•
•
-
•
plants
-
Eff-ct
-
-
-
'
'
-
-
- •
.
unspecified
.-
Study
Value
m0/k9>
4*i
•
•
•
' •
•
•
•
•
2
. •
D»«cfH»tfoft
X
-
-
'•
•
-
•
-
-
-
• - .
LOEC
-
&
-
-
•
-
-
•
-
•
' •-
•
-
-
-
-
'
•
-
:
Kabata-Pendias
and Pendias, 1984
a* cited in Will and
Suter, 1994.
'Benchmark Category, a = adequate, p * provisional, i « interim; ID - insufficient data; a (*) indicates that the benchmark value was an order
of magnitude or more above the NEL or UEL for other adverse effects.
August 1995
-------
APPENDIX B
Silver • 6
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconcentration factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values.
Insufficient data were identified to determine the whole-body BCF for silver in fish, aquatic
invertebrates, terrestrial vertebrates and earthworms. A whole plant BCF value of 4.0 E-01
was derived from Baes et al. (1983). For metals, empirical data were used to derive the BCF
for aboveground forage grasses and leafy vegetables. In particular, the uptake response slope
for forage grasses was used as the BCF for plants in the terrestrial ecosystem since most of
the representative plant-eating species feed on wild grasses.
Table 4. Biological Uptake Properties
receptor
fish
littoral
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF.BAF.t*
eSAF
-
-
-
BCF
ItoW-bM^W
vtoMxxtr
• -
•
•
whole-plant
, valu*
ID
ID '
ID
. ID
ID
4.0 E-01
_ ..
. _
•
•
'
Baesetal.. 1983
d = refers to dissolved surface water concentration
I = refers to total surface water concentration
ID = refers to insufficient data
August 1995
-------
APPENDIX B Silver-7
References -
AQUIRE (AOUatic Toxicity Information REtrieval Database), 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MR
Baes, C.F and R.D. Sharp. 1983. A proposal for estimation of soil leaching and leaching
constants for use in assessment models. J. Environ. Qual. 12(1): 17-28.
Carson, B.L. and I.C. Smith. 1975. Silver: An appraisal of environmental exposure. For
National Institute of Environmental Health Sciences, Research Triangle Park, NC.
Chapman, G.A., S. Ota, and F. Recht. 1980. Effects of water hardness on the toxiciry of
metals to Daphnia magna. U.S. EPA, Corvallis, OR:17p. As cited in AQUIRE (AOUatic
Toxicityjnformation REtrieval Database), 1995. Environmental Research Laboratory,
Office of Research and Development, U.S. Environmental Protection Agency, Duluth,
. MN. ' .
Clement Associates, Inc.. 1989. Draft: Toxicological Profile for Silver. Prepared for Agency
for Toxic Substances and Disease Registry (ATSDR), U.S. Public Health Service.
Connell, D.B., J.G.Sanders, G.F. Riedel, and G.R. Abbe. 1991. Pathways of silver uptake and
trophic transfer in estuarine organisms. Environ. Sci. Techno!. 1991,25, 921-924.
Davies, P.H., J.P Goettl, Jr. and J.R. Sinley. 1978. Toxicity of silver to rainbow trout (Salmo
gairdneri) Water Res. 12:113-117. As cited in LeBlane, G.A. 1984. Interspecies
relationships in acute toxicity of chemicals to aquatic organisms. Environmental
Toxicology and Chemistry, Vol.3, p47-60.
Day, W.A., J.S. Hunt and A.R. McGiven. 1976. Silver deposition in mouse glomeruli.
Pathology. 8:201-204. As cited in U.S EPA (Environmental Protection Agency). 1985.
Drinking Water Criteria Document for Silver. Prepared by the Environmental Criteria and
Assessment Office, Cincinnati, OH for the Office of Drinking Water.
Desquidt, J., P. Vasseur and J. Gromez-Potentier. 1974. Etude toxicologique experimentale de
quelques derive argentiques. 1. Localisation et elimination. (Experimental lexicological
study of some silver derivatives. 1. Localization and elimination.) Bull. Soc. Pharm. Lille.
1:23-35. As cited in U.S EPA (Environmental Protection Agency). 1985. Drinking Water
Criteria Document for Silver. Prepared by the Environmental Criteria and Assessment
Office, Cincinnati, OH for the Office of Drinking Water.
Elwell, W.S., J.W. Gorsuch, R.O.Kringle, K.A. Robillard, and R.C. Spiegel. 1986.
Simultaneous evaluation of the acute effects of chemicals on seven aquatic species.
.Environmental Tox and Chem, Vol.5, p831-840.
August 1995
-------
APPENDIX B Silver-8
57 FR 24152. June 5", 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg 3/4/day.
Hale, J.G. 1977. Toxicity of metal .mining wastes. Bull, of Environmental Contamination and
Toxicology 17(1).
Kabata-Pendias, A. and H. Pendias. 1984. Trace Elements in Soils and Plants. CRC Press,
Inc. Boca-Raton, Florida. As cited in Will, M.E and G.W. Suter II. 1994. Toxicological
Benchmarks for Screening of Potential Contaminants of Concern for Effects on Terrestrial
Plants: 1994 Revision. DE-AC05-84OR21400. Office of Environmental Restoration and
Waste Management, U.S. Department of Energy, Washington, DC.
LeBlanc, G.A. 1984. Interspecies relationships in acute toxicity of chemicals to aquatic
organisms. Environmental Toxicology and Chemistry, Vol.3, p47-60.
LeBlanc, G.A, J.D. Madstone, A.P. Paradice and B.F. Wilson. 1984. The influence of
speciation on the toxicity of silver to fathead minnow (Pimenphales promelas).
Environmental Tox and Chem., Vol.3,p 37-46.
Luckey, T.D. and B. Venugopal. Metal toxicity in mammals (1): Physiologic and chemical
basis for metal toxicity. Plenum Press, N.Y.
Mazbich, B.I. 1960. Some aspects of the pathogenesis of edema of the lungs due to silver
nitrate. Communication IE. The mechanism of death of experimental animals.
(Translated from Byull. Eksp. Biol. Med., 50(9):70-75. As cited in U.S EPA
(Environmental Protection Agency). 1985. Drinking Water Criteria Document for Silver.
Prepared by the Environmental Criteria and Assessment Office, Cincinnati, OH for the
Office of Drinking Water.
Mount, D.I and T.J. Norberg. 1984. A seven-day life cycle Cladoceran toxicity test.
Environ. Toxicol. Chem. 3(3):425-434. As cited in AQUIRE (AOUatic Toxicity
information REtrieval Database), 1995. Environmental Research Laboratory, Office of
Research and Development, U.S. Environmental Protection Agency, Duluth, MN.
Norberg-King, T.J. 1989. An evaluation of the fathead minnow seven day sub-chronic test for
estimating chronic toxicity. Environmental Tox and Chem, Vol.8, p 1075-1089.
Ridgway.L.P and D.A. Kamofsky. 1952. The effects of metals on the chick embryo: Toxicity
and production of abnormalities in development. Ann. N.Y. Acad. Sci. 55:203.
August 1995
-------
APPENDIX B Silver - 9
Rungby, J. and G. Danscher. 1984. Hypoactivity in silver exposed mice. Acta Pharmacol et
Toxicol 55:398-401. As cited in Clement Associates, Inc.. 1989. Draft: Toxicological
Profile for Silver. Prepared for Agency for Toxic Substances and Disease Registry
(ATSDR), U.S. Public Health Service.
Suter, G.W., and J.B. Mabrey. 1994. Toxicological benchmarks for screening potential
contaminants of concern for effects on aquatic biota: 1994 revision. ES/ER/TM-96/R1
Office of Environmental Restoration and Waste Management, U.S Department of Energy,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1980. Ambient Water Quality Criteria for
Silver. EPA-440/5-80-071. Office of Water Regulations and Standards, Criteria and
Standards Division, Washington, DC; Office of Research and Development,
Environmental Criteria and Assessment Office, Cincinnati, Ohio; Carcinogen Assessment
Group, Washington, DC, Environmental Research Laboratories, Corvallis, Oregon, Duluth
Minnesota, Gulf Breeze, Florida, Narrangasett, Rhode Island..
U.S EPA (Environmental Protection Agency). 1985. Drinking Water Criteria Document for
Silver. Prepared by the Environmental Criteria and Assessment Office, Cincinnati, OH for
the Office of Drinking Water.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume 1 and II. EPA 822/R-93-001a. Office of Water,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1993. Integrated Risk Information System.
February.
Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2): Chemical toxicity of metals
and metalloids. Plenum Press, N.Y., 1978.
Walker, F. 1971. Experimental argyria: A model for basement membrane studies. Br. J. Exp.
Pathol. 52(6): 589-593. As cited in U.S EPA (Environmental Protection Agency). 1985.
Drinking Water Criteria Document for Silver. Prepared by the Environmental Criteria and
Assessment Office, Cincinnati, OH for the Office of Drinking Water. Also cited in
Clement Associates, Inc.. 1989. Draft: Toxicological Profile for Silver. Prepared for
Agency for Toxic Substances and Disease Registry (ATSDR), U.S. Public Health
Service.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
August 1995
-------
Terrestrial Toxicity - Silver
CasNo.: 7440-22-4
Chemical Name
silver nitrate
silver nitrate
silver nitrate
silver nitrate
silver nitrate
silver nitrate
silver nitrate
silver nitrate
Species
dog
rats
rat
rat
mouse
mouse
rat
mice
Type of effect
circulatory and
pulmonary
effects
mortality
mortality
mortality
systemic
neurological
systemic
circulatory
Description
FEL
PEL
NOAEL
LOAEL -
NOAEL
LOAEL
PEL
PEL
Value
3.2
25.2
181.2
362.4
18.1
18.1
65
65
Units
mg silver
nitrate/kg
mg silver/kg
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mg/kg-d
mq/kg-d
Exposure Route
(oral, s.c., i.v.,
i.p.. injection)
.V.
i.p
oral (drinking
water)
oral (drinking
water)
oral (drinking
water)
oral (drinking
water)
oral (drinking
water)
oral (drinking
water)
Exposure
Duration /
Timing
<30min
24-48 hours
14-day
14-day
125-day
125-day
10 weeks
14 weeks
Reference
Mazhbich, 1960 as cited in
U.S. EPA, 1985
Dequidt et al., 1974 as cited
in U.S. EPA, 1985
Walker, 1971 as cited in
ATSDR, 1990
Walker, 1971 as cited in
ATSDR, 1990
Rungby & Danscher, 1984 as
cited in ATSDR, 1990
Rungby & Danscher, 1984 as
cited in ATSDR, 1990
Walker, 1971 as cited in U.S.
EPA, 1985
Day et at., 1976 as cited in
U.S. EPA, 1985
Comments
Dose killed 14 adult dogs less
than 30 minutes after lung edema,
arterial anoxia, and a drop in
blood pressure
Minimum i.p. dose of silver nitrate
resulting' within 24-48 hours.
Three of the twelve test species
died.
LOAEL is based on hypoactivity
silver deposits of basement
membrane
labeling of kidney capillary loops
with silver
NS = not specified
-------
Freshwater, ^xicity - Silver
Cas No. 744Q-22-4
Chemical Name
Silver
Silver
Silver
Silver
Silver
free Silver ion
free Silver ion
free Silver ion
free Silver ion
silver sulfide
silver thiosulfite
silver chloride
Silver
Species
Daphnia
pulex
Daphnia
maqna
Daphnia
magna
Fathead
minnow
Fathead
minnow
Daphnia
magna
Rainbow
trout
Bluegill
Fathead
minnow
Fathead
minnow
Fathead
minnow
Fathead
minnow
aquatic
organisms
Type of
effect
mortality
mortality
mortality
growth
mortality
mortality
mortality
mortality
mortality
mortality
mortality
mortality
NS
Description
LC50
EC50
EC50
LOEC
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
AWQC
Value
14
0.24
9.5
1.41
8.2
1.5
6.5
60
16
> 240,000
> 280,000
> 4,600
4.1
Units
ug/l
ug/l
ug/l
ug/l
UQ/1
ug/l
ug/l
ug/l
ug/l
ug/1
UQ/I
ug/1
ug/l .
Test Type
(static/ flow
through)
NS
NS
NS
Renewal
NS
NS
flow-through
NS
NS
NS
NS
NS
NS
Exposure
Duration /
Timing
2-day
2-day
2-day
7-day
7-day
48-hour
96-hour
96- hour
96-hour
96-hour
96-hour
96-hour
NS
Reference
Mount & Nor berg, 1984 as cited in
AQUIRE, 1995
Chapman et al., 1980 as cited in AQUIRE,
1995
Chapman et at., 1980 as Cited in AQUIRE,
1995
Norberg-King, 1989
Norberg-King, 1989
LeBlanc, 1984
Davies et al., 1978 as cited in LeBlanc et
al., 1984
LeBlanc, 1984
LeBlanc et al., 1984
LeBlanc et al., 1984
LeBlanc et al., 1984
LeBlanc etal, 1984
Suter& Mabrey, 1994
Comments
v
LOEC is based on reduced
growth in an Early Life-Stage test
water hardness < 1 00 mg/L
CaCO3
water hardness < 1 00 mg/L
CaCO3
AWQC value is a hardness
dependent criterion normalized to
100mq/l
NS = Not specified
-------
Freshwater Biological Uptake Measures - Silver
Cos No. 7440-22-4
Chemical
Name
Species
B-factor
(BCF. BAF,
BMP)
Value
Measured
or
Predicted
(m,p)
Units
Reference
Comments
-------
Terrestrial Biological l^.ake Measures - Silver
Cas No. 7440-22-04
Chemical
Name
silver
Spedes
whole-plant
B-factor
(BCF, BAF,
BMP)
BCF
Value
4.0 E-01
Measured
or
Predicted
(m-P)
m
units
(ug/g DW plant)/(ug/g
soil)
Reference
Baesetal, 1983
Comments
-------
APPENDIX B 2,3,7,8-TCDD - 1
Toxicological Profile for Selected Ecological Receptors
2,3,7,8-TCDD
CasNo.: 1746-01-6
Summary: This profile on 2,3,7,8-TCDD summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects reasonably assumed
to impact population sustainabiliiy. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
Kow between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents the most current information and may differ from the data presented in the
technical support document for the Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
1. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic planis, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Only one subchronic study documenting 2,3,7,8-TCDD exposure to mammalian
wildlife species was identified. Hochstein et al. (1986 as cited in Abt & Associates, 1993)
administered 2,3,7,8-TCDD dietary concentrations of 0, 0.001, 0.01, 0.1, 1.0, 10, and 100 ppb
10 mink for 125 days. While no significani adverse effects were observed on mink fed
dieiary concentrations of 0.1 ppb or less, mortality was noted in groups fed 1 and 10 ppb.
Several studies have documented subchronic and chronic exposure of 2,3,7,8-TCDD lo
laboratory animals. Khera and Ruddick (1972 as cited in U.S. EPA, 1993a) assessed the
postnatal effect of 2,3,7,8-TCDD on pregnant Wistar rats. In this experiment, rats were given
0, 0.125, 025, 0.5, or 1.0 ug TCDD/kg-day from days 6 through 15 of gestation. Dose related
decreases in ihe average litter size and pup weighi at birth were noted in all but the 0.125
August 1995
-------
APPENDIX B 2,3,73-TCDD - 2
ug/kg-day dose. Murray et al. (1979) exposed three generations of Sprague-Dawley rats to
diets containing 0, 0.001, 0.01, or 0.1 ug TCDD/kg-day. At the 0.01 ug/kg-day dose, Murray
et al. (1979) observed no effect on fertility among the f0 rats, but a significant reduction in
fertility was observed among the fl and f2 rats. Thus, through three successive generations,
the reproductive capacity of rats ingesting 2,3,7, 8-TCDD was clearly affected at dose levels of
0.01 and 0.1 ug/kg-day, but not at 0.001 ug/kg-day. Bowman et al. (1989a, 1989b) studied
the reproductive effects of Rhesus monkeys exposed to diets containing 5 ppt and 25 ppt
2,3,7,8-TCDD for 7 and 24 months. The female monkeys exposed to 25 ppt had a
significantly lower Index of Overall Reproductive Success (IORS), while the 5 ppt group did
not differ from the control. The 5 ppt was converted to a dose of 0.00013 ug/kg-day using
the study's daily allotment of 200 grams of monkey feed and the typical female monkey's
body weight outlined in Recommendations for and Documentation of Biological Values for
Use in Risk Assessment (U.S. EPA, 1988).
The study reported by Murray et al. (1979), in which three generations of Spraguer-Dawley
rats were exposed to 2,3,7,8-TCDD, was selected for developing a mammalian benchmark
value. This study was selected because it consists of a multi-generational exposure scenario
that demonstrates a clear dose-response for reproductive effects attributable to 2,3,7,8-TCDD.
The 125-day test performed by Hochstein et al. (1986 as cited in Abt & Associates, 1993)
was not considered as appropriate for deriving a benchmark since the study was subchronic,
rather than chronic and the perceived endpoints focus more on mortality than reproductive
effects. The Murray et al. (1979) study was chosen over the Khera and Ruddick (1972 as
cited in U.S EPA, 1993a) because of a lower reported NOAEL for rats. The reproduction
study by Bowman et al. (1989a, 1989b) on Rhesus monkeys (which produced a lower
NOAEL) was not selected because the Murray et al. (1979) study incorporated a
multigenerational exposure regime and contained stronger dose-response information.
The NOAEL of l.OE-6 mg/kg-d from the Murray et al. (1979) was scaled for species
representative of a freshwater ecosystem using a cross-species scaling algorithm adapted from
Opresko et al. (1994)
Benchmark = NOAEL, x
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BWt is the body weight of the test species. This is the default
methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57FR 24152). Since the
Murray et al..(1979) study documented reproductive effects from 2,3,7,8-TCDD exposure to
three generations of male and female rats, the mean male and female body weight for each
representative species was used in the scaling algorithm to obtain the lexicological
benchmarks.
August 1995
-------
APPENDIX B 2,3,7,8-TCDD - 3
Data were available oh the reproductive, developmental, and growth effects of 2,3,7,8-TCDD.
In addition, the data set contained studies which were conducted over chronic and subchronic
durations and during sensitive life stages. Most of the studies identified were conducted
using laboratory mammals and, as such, interspecies differences among wildlife were not
identifiable. Therefore, the data set does not support an uncertainty factor to account for
inter-species differences in toxicologieal sensitivity. The reproductive NOAEL selected from
Murray et al. (1979) was within an order of magnitude of the lowest identified NEL or LEL,
and therefore, the benchmarks developed for mammals representative of a freshwater
ecosystem were categorized as adequate.
Birds: In many field studies, reduced reproduction levels in avian species have been
correlated to 2,3,7,8-TCDD equivalents; however, the dose-response relationship specific to
2,3,7,8-TCDD itself cannot be determined from the effects of other contaminants. The only
identified research investigating the subchronic toxicity of 2,3,7,8-TCDD among avian species
was performed by Nosek et al. (1992). Ring-necked pheasants were dosed weekly by ip
injection for 10 weeks at an equivalent rate of 0.14, 0.014 and 0.0014 ug TCDD/kg-day
(weekly dose was divided by 7 for the equivalent daily dose). Cumulative egg production
was significantly reduced among pheasants exposed to 0.14 ug TCDD/kg-day, but not among
those pheasants exposed to the two lower doses.
The pheasant reproductive effect NOAEL of 0.014 ug/kg-day for 2,3,7,8-TCDD (Nosek et al.,
1992) was used in calculating avian wildlife benchmarks. The Nosek et al. (1992) study
demonstrates a clear dose-response to a critical reproductive endpoint and is based en an
exposure lasting more than 28 days. This study should be interpreted judiciously since it
involves an ip injection rather than an oral route of administration. Assuming 100%
absorption from ip injection, the ip exposure route may overestimate the absorption rate of
TCDD via oral ingestion by a factor of one to 5 depending upon diet composition (Abt &
Associates, 1993).
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric sealing for avian species were not identified. Thus, for avian species
representative of a freshwater ecosystem, the NOAEL of 0.014 ug/kg-d from Nosek et al.
(1992) was scaled using the cross-species scaling method of Opresko et al. (1994). Since the
Nosek et al. (1992) study documented reproductive effects from 2,3,7,8-TCDD exposure to
female pheasants, female body weights for each representative species were used in the
scaling algorithm to obtain the toxicologieal benchmarks. Although there is no formal
designation for benchmarks developed from ip exposure route studies, the benchmarks derived
from Nosek et al. (1992) were categorized as interim based on the absorption uncertainties
surrounding the intraperitoneal injection of TCDD to pheasants.
Fish and Aquatic Invertebrates: Since an AWQC for 2,3,7,8-TCDD was not available and a
Secondary Chronic Value (SCV) could not be calculated because of limited acute data, a
benchmark protective of fish and aquatic invertebrates was not established. However,
numerous fish studies documenting the effects of chronic 2,3,7,8-TCDD exposure were
identified. The rainbow trout is one of the most extensively studied aquatic organisms for
August 1995
-------
APPENDIX B 2,3,73-TCDD - 4
effects from 2,3,7,8-TCDD exposure. The lowest identified toxicity values for 2,3,7,8-TCDD
exposure to rainbow trout were a 4-day LC50 of 1.83 ng/1 (Bol et al. 1989 as cited in U.S.
EPA, 1993b) and a LOAEL of 0.038 ng/1 based on 45% mortality for a 28-day, flow-thru
water test (Mehrle et al., 1988). Based on the current data set (see master table), TCDD
appears highly toxic to aquatic organisms. This concern has prompted further research into
the aquatic data set of 2,3,7,8-TCDD and the applicability of LOEC/NOEC values toward
calculating a Final Chronic Value (FCV) or SCV with the eventual goal of establishing an
appropriate aquatic benchmark. • >
Aquatic plants: The lexicological benchmarks for aquatic plants were either: (1) a no
observed effects concentration (NOEQ or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (EC,,) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutwn).
Aquatic plant data was not identified for 2,3,7,8-TCDD and, therefore, no benchmark was
developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQp) method. The EQp method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^ to determine a chemical concentration that may
be present in the sediment while still protecting the benthic community (Stephan, 1993). The
EQp number is the best recommendation of a chemical concentration that may be present in
the sediment while still protecting the benthic community from harmful effects resulting from
possible chemical exposure. Since there is no AWQC, FCV, or SCV, the benchmark for the
benthic community was not calculated for 2,3,7,8-TCDD.
August 1995
-------
APPENDIX B
2,3,7,8-TCDD . 5
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
aptehe
mink
river otter
bald eagle
osprey
great blue
heron
mallard
lesser scaup
spottad
sandpiper
harrino gul
kmgfisher
tmUniMrtt
V*kM««k(K
4.3E-7(a)
2.6E-7 (a)
9.4E-6 (i)
1.2E-5(i)
1.1E-5(i)
1.3E-S(i)
1.5E-5(i)
2.9E-5 (i)
1.4E-5(i)
2.2E-5 (i)
<*» '
rat
rat
ring-necked
pheasant
ring -necked
pheasant
ring-necked
pheasant
ring -necked
pheasant
ring -necked
pheasant
ring -necked
pheasant
ring-necked
pheasant
ring-necked
pheasant
«**
rep
rep
rep
rep
rep
rep
rep
rep
rep
rap
*»*
VMM
«g«B4
1E-6
1E-6
1.4E-5
1.4E-S
1.4E-5
1.4E-5
1.4E-5
1.4E-5
1.4E-S
1.4E-5
, ~«~-
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
•F
-
-
-
-
-
•
-
•
e»m»ui>uMu»
-!
Murray et ej., 1979
Munay et al., 1979
Nosek et al., 1992
Noseketal.. 1992
Nosek et at., 1992
Nosek et al.. 1992
Noseketal., 1992
Nosek staJ., 1992
Nosek et el.. 1992
Noseketal., 1992
•Benchmark Category, a « adequate, p • provisional, i • interim; a - inolcales that the benchmark value was an order of
.magnitude or more above the NEL or LEL tor other adverse effects.)
August 1995
-------
APPENDIX B
2,3,7,8-TCDD - 6
Table 2. ToxicologicaJ Benchmarks for Representative Fish
Associated with Aquatic Ecosystem
IfeplMMWiV*
Si^efaa.
^^HfWV-
fish and aquatic
invertebrates
aquatic plants
benthic
community
Bwtxnirt
Y*k»
•of
10
ID
ID
-,»
-
-
- •
CflMt
•
'
_
-
.
-
(MBWiimmi
•
-
IL
'Benchmark Category, • * adequate, p « proviaional, i » interim; a "' indicate* that the benchmark value was an order of
magnitude or more above the NEL or LEL for other adverse effects.)
10 » Insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rational behind Toxicological benchmarks used to derive population
sustainability concentrations (PSC) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals and birds representing the generic terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, several toxicity
studies were identified that focused on the effects of 2,3,7,8-TCDD on mammals. Since no
additional studies for terrestrial mammals were found, the same surrogate study (Murray et
al., 1979) was used to calculate benchmark values for mammalian species representative of
terrestrial ecosystems. The NOAEL from the Murray et al. (1979) study was scaled for
species representative of a terrestrial ecosystem using a cross-species scaling algorithm
adapted from Opresko et al. (1994). Since the Murray et al. (1979) study documented
reproductive effects from 23.7,8-TCDO exposure to male and female rats, the mean of the
male and female body weights for each representative species was used in the scaling
algorithm to obtain the lexicological benchmarks. Based on the data set for 2,3,7,8-TCDD,
the benchmarks developed from the Murray et al. (1979) study were categorized as adequate,
as in the aquatic ecosystem..
Birds: No additional toxicity studies documenting terrestrial avian exposure to 2,3,7,8-TCDD
were identified. The Nosek et al. (1992) study, which documented a NOAEL for 2,3,7,8-
TCDD exposure to pheasants, was used as the benchmark value for avian species
representative of the terrestrial environment. Based on the avain dataset for 2,3,7,8-TCDD,
the benchmarks developed from the Nosek et al. (1992) study were categorized as interim.
August 1995
-------
APPENDIX B 2,3,73-TCDD * 7
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from percent
yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks, were selected
by rank ordering the LOEC values and then approximating the 10th percentile. If there were 10 or
fewer values for a chemical, the lowest LOEC was used. If there were more than 10 values, the 10th
percentile LOEC was used. Such LOECs applied to reductions in plant growth, yield reductions, or
other effects reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. However, terrestrial plant studies were not identified for 2,3,7,8-TCDD
and, as a result, a benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil community
were not identified.
August 1995
-------
APPENDIX B
2,3,7,8-TCDD - 8
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
$PS*hO
deer mouse
short-tailed
throw
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed
deer
red-tailed hawk
American
kestol
Northern
bobwhita
American robin
American
woodcock
plants
soil community
•mctaiMftt
VakM*«*9-4
1.2E-6(a)
1.2E-6(a)
9.2E-7 (a)
4.1E-7(a)
2.9E-7(a)
2.8E-7 (a)
1.4E-07(a)
1.3E-5(i)
2.3E-5 (i)
2.1E-50)
2.6E-5 (i)
2.0E-5 (i)
' ID
ID
«b*
•MtfM
rat
rat
rat
rat
rat
rat
rat
ring-mckad
pheasant
ring -nocked
pheasant
ring -nocked
pheasant
ring -necked
pheasant
ring -nocked
pheasant
Ettoct
rep
rep
rep
rep
rep
rep
rep
rep
f^>
rep
rep
rep
$tud)r
•V»M»
•9**^
1E-6
1E-6
1E-6
1E-6
1E-6
1E-6
1E-6
1.4E-5
V4E-5
1.4E-5
1.4E-5
1.4E-5
J^B^aAV^a^^e&
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
**
-
•
•
,
•
' •
-
-
-
-
-
&WM)t9MNWt -• ^
Murray etal., 1979
Murray eta).. 1979
Murray el a)., 1979
Murray etai:, 1979
Murray et a!., 1979
Murray etal.. 1979
Murray et a).. 1979
Nosek at al., 1992
Nosekotal.. 1992
Nosek etai., 1992
Nosek et el., 1992
Nosek eta).. 1992
•Benchmark Category, a - adequate, p • provisional, i «interim; a "' indicates that the benchmark value was an order oi
magnitude or more above the NEL or LEL for other adverse effects.
ID * Insufficient Data
August 1995
-------
APPENDIX B 2,3,7,8-TCDD - 9
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log Kow values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log Kow values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context for interpreting the biological uptake values presented in
Table 4.
Because the log Kow for TCDD is above 6.5 (i.e., 7.04), the Thomann (1989) and Thomann et
al., (1992) models were not used to estimate bioaccumulation factors. The lipid-based
bioaccumulation factors for fish and invertebrates in the limnetic ecosystem were taken from
the Great Lakes Water Quality Initiative Tec,uiical Support Document for the Procedure to
Determine Bioaccumulation Factors - My 1994 (U.S. EPA, 1994b). The document indicated
that the basis for the TCDD BAF,d s was a memorandum from P.M. Cook to C.E. Stephan
(July, 1994) that, in particular, recommended that the BAF,d for trophic level 3 be a factor of
two higher than that for trophic level 4. These values are consistent with the work previously
presented by Cook (U.S. EPA, 1993i) on bioaccumulation of TCDD in fish and represents the
current state-of-the-science at the Agency. However, for extremely hydrophobic constituents,
the Agency has stated that reliable measurements of ambient water concentrations (especially
dissolved concentrations) are not available and that accumulation of these constituents in fish
or other aquatic organisms cannot be referenced to a water concentration as required for a
BCF or BAF (U.S. EPA, 1993i). Fortunately, extremely hydrophobic constituents can be
measured in sediments and aquatic life and, because these chemicals tend to partition to lipids
and organic carbon, a biological uptak£ factor that reflects the relationship between sediment
concentrations and organism concentrations may be more appropropriate. Consequently, the
BSAF is the preferred metric for accumulation in the littoral aquatic ecosystem for extremely
hydrophobic chemicals (e.g., chemicals with > log K,,w of - 6.5). The biota-sediment
accumulation factor (BSAF) in [mg TCDD/kg LP]/[mg TCDD/kg sediment OC] for trophic
level 4 fish was supplied by the U.S. EPA ORD Exposure Assessment Group in a
memorandum to Addressees by Matthew Lorber (September, 1994). This memorandum
updates the Addendum to the Methodology for Assessing Health Risks Associated with
Indirect Exposure to Combustor Emissions (U.S. EPA, 1993a) and other EPA documents
. involving risk assessment of 2,3,7,8-TCDD. As with the BAF,ds, this recommendation
August 1995
-------
APPENDIX B 2,3,7,8-TCDD - 10
represents the current" state-of-the-science at the Agency. The BSAF for trophic level 3 fish
was calculated as the geometric mean of BSAFs for "smaller" fish presented by Cook (U.S.
EPA, 19931) such as perch, smelt, and sculpin. Although these fish may not be strictly
regarded as trophic level 3 species, they are reasonable species to represent fish eaten by
larger piscivorous fish. The BSAF for trophic level 2 invertebrates was approximated by
using a laboratory mean BSAF for the sandworm (Rubinstein et al., 1983 as cited in U.S.
EPA, 1993i). Although selecting a single species value for this trophic level is associated
with greater uncertainty than a geometric mean of multiple species (under different
conditions), the sandworm is an appropriate species to represent sediment dwellers likely to
be eaten by the ecological receptors.
The bioconcentration factor (BCF/) for fish was estimated by calculating the geometric mean
of measured values presented in the Interim Report on Data and Methods for Assessment of
2J,7,8-Tetrachlorddibenzo-p-dioxin Risks to Aquatic Life and Associated Wildlife (U.S. EPA,
1993i). The geometric mean BCF/ of 515,251 is approximately a factor of 3 higher than the
BCF/ estimated from the BCF/1 <• log Kow relationship and adjusted for the dissolved fraction
Od) as defined in Equation 6-21. It is somewhat surprising that the BCF/ based on measured
values is higher than the predicted value since TCDD has been shown to be metabolized in
fish (U.S. EPA, 1993i). However, the discrepancy is likely the result of uncertainties
surrounding* the most appropriate log Kow to use for TCDD mixtures as well as differences in
water conditions (e.g., total suspend solids) in different studies. Nevertheless, the difference
between the two values was considered insignificant given the inherent uncertainties in BCF
measurement and modeling techniques, and t^ slightly more conservative BCF/ of 500,000
(rounded) was selected as the bioconcentration factor in general fish.
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of a number of
measured values with sources shown in Table 4 (see master table). For terrestrial
invertebrates and earthworms, high-end bioconcentration factors were selected from the
Revision of Assessment of Risks to Terrestrial Wildlife from TCDD'and TCDF in Pulp and
Paper Sludge (Abt, 1993). For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described in Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
2,3,7,8-TCDD - 11
Table 4. Biological Uptake Properties
eootoflleal
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCfvtAF,*
83AF
BAF
BAF
BCF
BSAF
BSAF
BSAF
BAF
BCF
BCF
BCF
HpW-bawdor
irVtiota1 pody
lipid
lipid
lipid
}
lipid
lipid
lipid
whole-body
whole-body
whole-booy
whole-plant
Value
7,850,000 (d)
15,700.000 (d)
500,000 (t)
0.067
0.068
0.48
7.2
1.3
9.1
0.0033
0
eoutc*
measured value from Cook, 1M4
as cited in U.S. EPA, 1M4b
measured value' from Cook, 1904
as cited in U.S. EPA, 1994b
approximate geomeSx mean of
measured value* in U.S. EPA,
19031
recommendation by the U.S.
EPA ORO, 1905
geometric mean of BSAF values
in U.S. EPA, 1993 tor 'smeJIer*
trophic level 3 fish
BSAF value in U.S. EPA. 1903i
geometic mean of measured
values (e.g., Garten and
Trabafea. 1983; Abt Associates.
1993)
high-end measured value from
Abt Associates. 1993
high-end measured value from
Abt Associates, 1903
U.S. EPA, 1992e (does not
include ajr-to-piant)
d - refers to dissolved surface water concentration
t » refers to total surface water concentration
August 1995
-------
APPENDIX B 2,3,7,8-TCDD . 12
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Biological Values for Use in Risk Assessment. Cincinnati, Ohio. PB 88-179874. 21 pp.
U.S. Environmental Protection Agency. 1990. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment. Washington, D.C. January. As cited in Pierson,
T.K., A.E. Crook, S.M. Beaulieu, P.N. Graham, N.B. Jones, A.M. Reynolds, and G.P.
Vegh. 1994. Development of Human Health Based Exit Criteria for the Hazardous
Waste Identification Project. Phase III Analysis. February.
August 1995
-------
APPENDIX B 2,3,73-TCDD . i8
U.S. Environmental Protection Agency. 1992a. 304(A) Criteria and Related Information for
Toxic Pollutants. Water Management Division - Region IV.
U.S. Environmental Protection Agency. 1993i. Interim Report on Data and Methods for
Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioxin Risks to Aquatic Life and Associated
Wildlife. EPA/600/R-93/055. Office of Research and Development, Washington, DC.
U.S. Environmental Protection Agency. 1994a. Estimating Exposure to Dioxin-Like
Compounds. Volume III: Site-Specific Assessment Procedures. EPA/600/6-88/005Cc.,
Office of Research and Development, Washington, DC.
U.S. Environmental Protection Agency. 1994b. Great Lakes water Quality Initiative
Technical Support Document for the Procedure to Determine Bioaccumulation Factors.
EPA-822-R-94-002.
Veith, G.D., D.L. Defor, B.V. BergstedL (1979). Measuring and Estimating the
Bioconcentration Factor of Chemicals in Fish. /. Fish Res. Board Can. 26, 1040-1048. As
cited in Mackay, Donald, Wan Ying Shiu, and Kuo Ching Ma. 1992. Illustrated
Handbook of Physical-Chemical Properties and Environmental Fate for Organic
Chemicals. Vol. H, pp. 400-409.
Walker, M:K., J.M. Spitsbergen, J.R. Olson and R.E. Peterson. 1991. 2,3,7,8-
Tetrachlorodibenzo-p-dioxin Toxicity Du~' :g Early Life Stage Development of Lake Trout
(Salvelinus namaycush). Can. J. Fish Aquat. Sci. 48:875-883. As cited in U.S.
Environmental Protection Agency. 1992b. Chapter 5. Reproductive and Developmental
Toxicity (Review Draft). Office of Research and Development, Washington, DC.
EPA/600/AP-92/001e
Walton, B.T. and N.T. Edwards. 1986. Accumulation of Organic Waste Constituents in
Terrestrial Biota. Water Resources Symposium - No. 13, Land Treatment: A Hazardous
Waste Management Alternative, pp. 73-86.
Will, M.E. and G.W. Suter, 1994. Toxicolpgical Benchmarks for Screening Potential
Contaminants of Concern for Effets on Terrestrial Plants: 1994 Revision. ES/ER/TM-
85/R1. Prepared for U.S. Department of Energy.
Young, A.L., C.E. Thalken, and D.D. Harrison. 1981. Persistence, Bioaccumulation, and
Toxicology of TCDD in an Ecosystem Treated with Massive Quantities of 2,4,5-T
Herbicide. Proc. Wast. Soc. Weed Sci., Vol. 34, pp. 70-77.
August 1995
-------
Terrestrial Toxic. , - 2,3,7,8-TCDD
Cos No.: 1746-01-6
Chemical Name
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3.7,8-TCDD
2,3,7,8-TCDD
2,3,7.8-TCDD
2,3,7,8-TCDD
2 37.8'TCDD
Species
rat
.
rat
rat
mink
(adult)
Rhesus
monkey
Rhesus
monkey
Rhesus
monkey
Endpoint
rep
rep
, f
NS
mortality
rep
rep
rep
Description
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
LOAEL
NOAEL
•
Value
0.125
0.001
0.001
0.0055
0.00021
0.0017
0.0095
Units
ug/Kg-day
ug/kg-dav
ug/kg-day
ug/kg-body wt
ug/kg-day
ug/kg
ug/kQ-body wt.
Exposure
Route (oral,
B.C., I.V., l.p.,
inlectton)
NS
oral
dietary study
dietary study
oral
diet
diet
Exposure
Duration/
Timing
days 6 - 1 5 oJ
gestation
at least 90
days prior to
gestation and
throughout
gestation
2-year
125 day
7 and 24
months
7 to 24
months
3 x weekly (or
3 weeks
Reference .
Khera & Ruddick.
1972 as cited in
U.S. EPA, 1993a
Murray et at.. 1979
Kociba et al., 1978
Hochstein el al..
1986 as cited in Abt
& Associates, 1993
Bowman el al.,
19898, 1989D
Eisler, 1986
Eisler, 1986
Comments
Dose-related decreases in the
average litter size and pup weight at
birth were noted in all but the 0.125
ug/kg-day dose level.
Three generation study. At 0.01
ug.kg-day dose, significant reduction
in fertility was observed among the
F1 and F2 rats. No difference was
observed between the fertility of the
0.001 ug/kg-day rats and the
controls.
Administered dietary cone, of 0,
0.001. 0.01. 0.1. 1.0, 10, and 100
ppb. No significant adverse effects
observed on mink fed a dietary
concentration of 0.1 ppb or less -
mortality noted in groups fed 1 and
10 ppb.
The 25 ppt group of mothers had a
significantly lower Index of Overall
Reproductive Success, while the 5
ppt group did not differ from the
control.
Abortion and weight losses were
reported
No adverse effects on reproduction
-------
Terrestrial Toxlclty - 2,3,7.8-TCDD
Cos No.: 1746-01-6
Chemical Name
2,3.7,8-TCDD
2,3,7.8-TCDD
2.3,7.8-TCDD
2,3,7,8-TCDD
2.3,7,8-TCDD
2,3.7.8-TCDD
2,3,7,8-TCDD
2,3.7.8-TCDD
2,3.7.8-TCDD
237 8-TCDD
Spedes
rat
mouse
rabbit
hamster
guinea pig
guinea pig
guinea pig
guinea pig
mink
(adult)
doq
Endpoint
acute
acute
acute
acute
.t'
acute
acute
acute
acute
acute
acute
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
22-45
114-284
115
1157-5051
0.6
0.6
2
0.8
4.2
100-200
Units
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ug/kg.
ug/kg.
ug/kg:
ug/kg.
ug/kg-body wt.
uo/kg-bodv wt.
Exposure
Route (oral,
S.C.. I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral
dietary study
oral
oral
Exposure
Duration/
Timing
NS
NS
NS
NS
single dose
single dose
single dose
NS
single dose
NS
Reference
Kociba & Schwetz,
I982a.bascitedin
Eisler, 1986
Kociba & Schwetz,
1982a,bascitedin
Eisler, 1986
Olson et al..
1980a,bascitedin
Eisler, 1986
Kociba & Schwetz.
1982a,bascitedin
Eisler, 1986
Schwetz etal.,
1973 as cited in Abt
& Associates. 1993
Hariess et al., as
cited in Eisler. 1986
Kociba & Schwetz,
19B2a,bascitedin
Eisler. 1986
DeCaprio et al.,
1986 as cited in Abt
& Associates. 1993
Hochstein et al.,
1988
Kociba & Schwetz.
I982a,bascitedin
Eisler, 1986
Comments
•
Adult male mink were administered a
single oral dose and the mink were
observed for 28 days.. A 28-day
LD50 value of 4.2 ug/kg body weight
was calculated.
.
i
-------
Terrestrial Toxk , - 2,3,7,8-TCDD
Cos No.: 1746-01-6
Chemical Namd
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3,7,8-TCDD
2,3.7,8-TCDD
2,3.7.8-TCDD
2,3,7,8-TCDD
2.3 7,8-TCDD
Spedes
Rhesus
monkey
northern
bobwhite
quail
ringed
turtle dove
mallard
domestic
chicken
ring-
necked
pheasant
(embryo)
ring-
necked
pheasant
(embryo)
guinea pig
CF-1 mice
Endpoint
acute
acute
acute
atiute
acute
acute
^
acute
subchronic
terat
Description
LD50
LD50
LD50
LD50
LD50
LD50
LD50
NOAEL
NOAEL
Value
<70
15
>810
i
>108
25-50
1,354
2,182
0.65
0.1
Units
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ug/kg-body wt.
ug/kg-bo^y wt.
pg TCDD/g egg
pg TCDD/g egg
ng/kg-day
ug/kg-day
Exposure
Route (oral,
S.C.. I.V., i.p.,
injection)
?
oral
oral
oral
oral
oral
ovo
injections
ovo
injections
dietary study
oral qavaae
Exposure
Duration /
Timing
NS
NS
NS
NS
NS
single .
injection
single
injection
7
90 day
days 6- 15 of
gestation
Reference
Olson et al.,
1980a,bascitedin
Eisler, 1986
Hudson etal.. 1984
as cited in Eisler,
1986
Hudson etal.. 1984
as cited in Eisler,
1986
Hudson etal., 1984
as cited in Eisler.
1986
Kociba A Schwetz.
1982a.bascitedin
Eisler. 1986
Noseketal., 1993
.
Noseketal., 1993
DeCaprio et al ,
1 986 as cited in Abt
& Associates, 1993
Smith el al., 1976
Comments
Fertilized eggs were injected into the
albumin with vehicle or graded doses
Ol2.3.7,8-TCDD(001,0.1. 1. 10.
100. 1.000, 10.000, or 100.000 pg
TCDD/g egg) on day 0 toxicity was
assessed in 1-d hatchlings and 28-
day chicks.
Fertilized eggs were injected into the
yolk with vehicle or graded doses of
2.3.7.8-TCDD (0.01, 0.1, 1. 10, 100.
1,000. 10.000, or 100,000 pg TCDD/J
egg) on day 0 toxicity was assessed
n 1-d hatchlings and 28-day chicks.
Doses were 0, 0.001 , 0.01 , 0. 1 . 1 .
and 3 ug/kg-day Cleft palate was
ound at 1 .0 uo/kq-dav dose
-------
Terrestrial Toxiclry - 2.3,7,8-TCDD
Cos No.: 1746-01-6
Chemical Name
.
2,3,7,8-TCDD
2 3.7 8-TCDD
Species
while
leghorn
chicken
ring-
necked
pheasant
,
Endpoint
subchronic
rep
Description
NOAEL
NOAEL
Value
01
0:014
Units
ug/kg-day
-
uo/ka-dav
Exposure
Route (oral.
s.c.. i.v., l.p.,
infection)
oral
intubation
i.p.
Exposure
Duration /
Timing
20 - 21 day
i.p. injection
once a week
for 10 weeks
Reference
Schwetz elal.,
1973 as cited in Abt
& Associates, 1993
Noseketal.. 1992
Comments
Dose was administered in a corn
oil/acetone vehicle to 3-day old
chickens for 21 days - no evidence of
chick edema or gross lesions were
found.
NOAEL was converted to a daily'
dose from a weekly dose. TCDD
dose of 0.14 ug/kg-day reduced egg
production and eqq hatchabilitv
NS = not specified
I
-------
freshwater Toxk.../ - 2.3,7,8-TCDD
Cos No.: 1746-01-6
Chemical Name
2.3.7.8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
2.3.7.8-TCDD
2.3,7,8-TCDD
2.3,7.8-TCDD
2,3.7.8-TCDD
Species
rainbow (rout
rainbow trout
$
rainbow trout
lake trout
eggs
yellow perch
coho salmon
daphnia
magna
rainbow trout
fathead
minnow
Endoolnt
growth
growth
retardation
toxic effects
sac try
mortality
toxic effects
reduced
survival and
growth
acute .
acute
acute
Description
NOEC
LOEC
NOEC
CV
NOEC
NOEC
NOEC
LC50
LC50
Value
<38
0.1
494
43
494
0.56
1,030
1.83
1.7
Units
pg/l
ng/l
pg TCDD/g diet
•
pg TCDD/g «gg
pg TCDD/g diet
'
ng/L
ng/l
ng/l
noA
Test Type
(static/ flow
through)
flow-through
NS
flow through
NS
flow through
static
water
renewal
water
renewal
water
renewal
Exposure
Duration /
Timing
28 day
exposure. 28
day
depuration
period
96-hour
exposure.
24-week
observatrion
13 weeks
NS
13 weeks
12-hour
exposure,
114-day
observation
perioto
2-day
4-day
28-day
Reference
Mehrteelal., 1988
Hekter, 1981
Kleeman et al..
1986a
Walker etal.,
1991 as cited in
U.S. EPA. 1992b
Kleeman et al.,
I986b
Miller etal.. 1979
as cited in U.S.
EPA, 1993c
Adams etal., 1986
Boletal., 1989 as
cited in U.S. EPA.
19931
Adams etal.. 1986
Comments
The NOEC ot TCDD on growth
during the exposure and
depuration phases was less than
the lowest exposure '
concentration ol 38 pg/l
0.1 ng/l resulted in a significant
growth retardation (or 72 days.
Fingerting trout weighed 7 - 14 g
when dietary exposure to TCDD
began. Trout diet (ed at 3% body
wt./day
Fingerting perch weighed 5 - 1 0 g
when dietary exposure to TCDD
began. Perch diet fed at 3%
body wt /day
Cone, of 0.56 ng/l had no effect
on food consumption . weight
gain, or survival but 5.6 ng/l
reduced survival and growth 1 14
days after an exposure ol 12
hours.
48 hour exposure, followed by 7-
day observation period.
NS = Not specified
-------
Freshwater Biological Uptake Measures - 2.3.7,8-TCDD
Cos No.: 1746-01-6
Chemical Nam*
2,3,7,8-TCDD
2.3,7,8-TCDD
2,3.7,8-TCDD
2,3.7,8-TCDD
2,3,7,8-TCDD
2.3.7,8-TCDD
2.3,7,8-TCDD
2.3.7,8-TCDD
2,3,7,8-TCDD
2,3.7,8-TCDD
2.3,7.8-TCDD
Species
rainbow trout
rainbow trout
fish
fathead minnow
,1
fathead minnow
carp
fish
fish
fish
fish
lish
B-factor
(BCF. BAF.
BMR
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
Valu*
39,000
116.000
9,270
113,000
510,000-
837,000
733,000
5,370
35.481
5,000
13,158
33.884
Measured
or
Predicted
(m.p)
m
m
P
m
m
m
m
m
m
P
P
Units
NS
NS
NS
NS
NS
NS
NS
NS
L/Kg
L/kg
NS
Reference
Mehrteetal, 1988
Cook et aj., 1991
Branson et al.,
1985
Adams et al., 1986
Cook et at., 1991
Cook el al., 1991
Kenaga & Goring.
1980 as cited in
Mackay et al., 1992
Kenaga & Goring,
1980 as cited in
Mackay et al'.. 1992
U.S. EPA. 1992a
Slephan, 1993
Neelyelal . 1974
as cited in Mackay
elal , 1992
Comments
Using Branson et al. ( 1 985) study,
Cook assumed 8% lipid based on fish
species, size, and age
Assuming 7% lipid based on fish
species, size, and age
19%lipids.
9% lipids
Flowing water test conditions.
Static ecosystem conditions
3% lipid
1% lipid
-------
Freshwater Biological Uph. > Measures - 2,3,7,8-TCDD
Cos No.: 1746-01-6
Chemical Name
2.3.7.8-TCDD
2.3.7,8-TCDD
2.3,7.8-TCDD
2.3,7,8-TCDD
2.3.7.8-TCDD
2.3.7.8-TCDD
Species
fish
flsh
fish
fish
white sucker
fish
B-factor
(BCF. BAF.
BMP)
BCF
BCF
BCF
BCF
BAF
BAF
Value
33,113
1.995
1.047
239,883
37,160
60.000
Measured
or'
Predicted
(m.p)
P
P
P
P
m
p
Units
NS
NS
HO/I-
NS
(pg TCDD/g
fish)/(pg
TC 0/g
water)
NS
Reference
Viethetal.. 1979
as cited in Mackay
el al., 1992
Banerjee et al ,
1980 as died in
Mackayetal., 1992
Garten & Trabalka.
1983
Chlouetal, 1977
as cited Branson et
al., 1985
Frakesetal., 1993
Cook; 1992 as
cited In Stephen,
1993
Comments
Microcosm condition.
4 different rivers sampled. Geometric
mean of all the BAFs combined =
37.160
Normalized to 5% lipid. BAF
calculated for fish jfophic levels 3 and
4. 7
NS = not specified
-------
Terrestrial Biological Uptake Measures - 2.3,7.8-TCDD
Cos No.: 1746-01-6
Chemical Name
2,3.7,8-TCDD
2.3.7.8-TCDD
2.3.7.8-TCDD
2,3,7.8-TCDD
2.3.7.8-TCDD
2.3.7.8-TCDD
2.3,7.8-TCDD
2.3,7.8-TCDD
2.3.7,8-TCDD
2.3.7.8-TCDD
237 8-TCDD
Species
plant
earthworms
rats
cattle
Rhesus monkey
earthworms
Insects and other
invertebrates
mammals (and
other vertebrates)
earthworms
rodents
cow
B-factor
(BCF. BAf .
BMF)
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BCF
BAF
BAF
BAF
Value
5.60E-03
15
3.7 - 24.5
24.8
24-40
10
1.25
1.3
5
0.71
3.5
Measured
or
Predicted
(m.p)
p
m
m
P
P
P
P
P
m
P
p
Units
NS
(ua/ka)/(ua/ko.)
NS
NS
NS
NS
NS
NS
(ug TCDD/g
worm)/(ug
TCDD/g soil)
(mg/kg of
fat)/(mg/kg of
diet)
(mg/kg of
fat)/(mg/kg of
diet)
Reference
U.S. EPA. 1990e
Martinucci et al.
1983 as cited In
Watton & Edwards.
1986
Kocibaetal.. 1978
as cited In Geyer
etal.. 1986
Jensen etal.. 1981
as cited In Geyer
etal.. 1986
Bowman et al..
1985 as cited In
Geyer etal.. 1986
Abt & Associates.
1993.
Abt & Associates.
1993.
Abt & Associates.
1993.
Relnecke & Nash.
1984
Garten &
Trabalka. 1983
Garten &
Trabalka, 1983
Comments
Plant uptake from soil pertains to
leafy vegetables.
BCF values In adipose tissue of rats.
Percent lipld was not specified
BCF values in fat of monkeys.
High-end exposure estimate
a'
e estimate
High-end exposure estimate
In soils containing 0.05 ug/g
earthworms accumulated TCDD
up to 5 times the original soil
concentration within 7 days.
% lipld was not specified in study.
% lipid was not specified in study.
NS = not specified
-------
APPENDIX B Toxaphene -1
Toxicological Profile for Selected Ecological Receptors
Toxaphene
CasNo.: 8001-35-2
Summary: This profile on loxaphene summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals, daphnids and fish, aquatic plants and benthic organisms
representing the generic freshwater ecosystem and birds, mammals, .plants, and soil
invertebrates in the generic terrestrial ecosystem. Toxicological benchmarks for birds and
mammals were derived for developmental, reproductive or other effects ^reasonably assumed
to impact population sustainability. Benchmarks for daphnids, benthic organisms, and fish
were generally adopted from existing regulatory benchmarks (i.e., Ambient Water Quality
Criteria). Bioconcentration factors (BCFs), bioaccumulation factors (BAFs) and, if available,
biomagnification factors (BMFs) are also summarized for the ecological receptors, although
some BAFs for the freshwater ecosystem were calculated for organic constituents with log
KOW between 4 and 6.5. For the terrestrial ecosystem, these biological uptake measures also
include terrestrial vertebrates and invertebrates (e.g., earthworms). The entire lexicological
data base compiled during this effort is presented at the end of this profile. This profile
represents ihe most current information and may differ from the data presented in the
technical support document for Ihe Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
\-
This section presenls Ihe rationale behind lexicological benchmarks used lo derive protective
media concentrations (C^) for the generic freshwater ecosystem. Table 1 conlains
benchmarks for mammals and birds associated with the freshwaier ecosystem and Table 2
conlains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found for mammalian wildlife in
which dose-response data were reported. However,, several chronic and subchronic toxicity
studies involving toxaphene have been conducted using laboratory rats and mice. A
August 1995
-------
APPENDIX B Toxaphene-2
reproductive study was identified in which groups of 4 male and 14 female Swiss white mice
were administered a dietary concentration of 25 ppm of toxaphene for six generations
(Keplinger et'al., 1970). Keplinger et al., (1970) observed litter size, survival rate, fetal
mortality, body weight, lactation, and reproduction and recorded a NOAEL of 25 ppm. In a
similar study, reproductive toxicity was also observed in 8 male and 16 female rats (Kennedy
et al.,,1973) fed diets containing 25 or 100 ppm toxaphene. In this three generation study,
Kennedy et al., (1973) observed growth, mortality, organ weights, litter size, pup survival or
weanling body weights and reported a NOAEL of 100 ppm for reproductive effects. Allen et
al. (1983) fed toxaphene to female weanling Swiss-Webster mice at doses of 10, 100, and 200
ppm for 8 weeks. In the fetuses, the 100 ppm dose caused a slight suppression of the DTK
response, significant impairment of humoral immunity, and almost complete inhibition of the
macrophage phagocytic ability. Chernoff and Carver, (1976) observed fetotoxicity in
random-bred albino CD-I mice and CD rats fed diets containing gavage doses of 15, 25, and
35 mg/kg-day of toxaphene (in corn oil) during days 7-16 of gestatipn. For the CD-I mice
Chernoff and Carver observed no dose-related effects on fetal mortality, fetal weight, number
of caudal or sternal ossification centers, or incidence of supernumerary ribs at any of the dose
levels. However, Chernoff and Carver noted a reduction in the average number of sternal and
caudal ossification centers in the CD rats and reported a LOAEL of 15 mg/kg-day for
developmental toxicity.
The LOAEL of 15 mg/kg-day in the Chernoff and Carver (1976) study was chosen to derive
the toxicological benchmark because: (1) dietary exposures were administered via oral
ingestion during a critical life-stage period, (2) it focused on developmental toxicity as a
critical endpoint, and (3) the study contained dose-response information. The Allen et al.
(1983) study, which reported a lower NOAEL (100 ppm converted 1.9 mg/kg-day) than the
selected representative study, was not selected because the reproductive effects were
considered not significant enough to impair population sustainability. The studies by
Keplinger et al., (1970) and Kennedy et al. (1973) were not chosen for the development of a
toxicological benchmark because they lacked dose-response data. Nevertheless, these studies
illustrate the dose ranges at which toxicity occurs.
The study value from Chernoff and Carver (1976) was divided by 10 to provide for a
LOAEL-to-NOAEL safety factor. This value was then scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994) where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body
weight of the wildlife species, and BW, is the body weight of the test species. This is the
default- methodology EPA proposed for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57111 24152). Since
August 1995
-------
APPENDIX B Toxaphene-3
Chernoff and Carver (1976) documented reproductive effects from toxaphene exposure to
female rats, female body weights for each representative species were used in the scaling
algorithm to obtain the toxicological benchmarks.
f bw T
Benchmarkw = NOAEL, x L
' (bwj
Data were available on the reproductive and developmental effects of toxaphene, as well as
chronic survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations and during sensitive life stages. The studies identified were
not conducted using a range of wildlife species and therefore, inter-species toxicity
differences were not identifiable. There were several study values in the data set which were
more than a magnitude lower than equal to the benchmark value. These values corresponded ,
to effects on behavioral, neurologic, and immunologic endpoints. Based on the data set for
toxaphene and because the benchmark is based on a LOAEL/10, the benchmarks developed
from Chernoff and Carver (1979) were categorized as provisional, with a "*" to indicate that
adverse effects may occur at the benchmark level.
Birds: No subchronic or chronic studies, with adequate dose-response regimes, were
identified for toxaphene exposure to the representative avian species. However, subchronic
and chronic toxicity studies involving toxaphene exposure to chickens, black ducks, and
pheasants have been conducted. In a subchronic study by Bush et al. (1977) female white
leghorn chickens, from 1 day-old, were fed a dietary concentrations of 0.5, 5, 50, and 100
ppm toxaphene for a period of 30 weeks. Birds fed 5, 50, and 100 ppm toxaphene exhibited
sternal or keel deformation at 30 weeks of age. Histopathological examination revealed renal
lesions in birds fed at the 50 and 100 ppm levels. Based on the reference body weight and a
food ingestion equation representative of chickens (U.S. EPA, 1988), a time-weighted
average NOAEL of 0.038 mg/kg-day was calculated. In another study involving avian
exposure to toxaphene, Genelly and Rudd (1958) fed pheasants 100 and 300 ppm toxaphene
in their diets for 2 to 3 months. The 300 ppm dose corresponded to a decrease in egg laying,
hatchability, food intake, and weight gain. Both dose levels in this study (Genelly and Rudd,
1958) caused greater mortality in young pheasants during the first 2 weeks after hatching than
was observed in the control group. Heinz, and Finley (1978) observed no change in avoidance
behavior (life-threatening if interrupted) in American black ducks fed dietary concentrations
of 10 or 50 mg/kg-diet of toxaphene. In a reproductive study (Haseltine et al., 1980),
American black ducks were fed 10 or 50 mg/kg-diet of toxaphene over a 19-month period.
August 1995
-------
APPENDIX B , Toxaphene-4
Haseltine et al., (1980) observed survival, egg production, fertility, hatchability, eggshell.
thickness, or growth or survival of young and recorded a NOAEL of 50 mg/kg-diet.
The NOAEL reported by Bush et al. (1977) was selected as the toxicological benchmark
representative of avian species because it was the lowest toxicity value in the dataset, had
sufficient dose-response data, and focused on developmental toxicity during a critical life-'
stage period. The study by Genelly and Rudd (1958) on pheasants was not selected as a
benchmark derivation based on a dual concern that the study was outdated and that the dose-
response correlation was not as strong as recorded by Bush et al. (1977). The studies by
Pollock and Kilgore (1978) on ring-necked pheasants and Heinz and Finley (1978) on
American black ducks were not selected for the derivation of the benchmark because of the
lack of dose-response information.
The principles for allometric scaling were assumed to apply to birds, although specific studies
supporting allometric scaling for avian species were not identified. Thus, for the avian
species representative of a freshwater ecosystem, the NOAEL of 0.038 mg/kg-day from the
Bush et al. (1977) study was scaled using the cross-species scaling method of Opresko el: al.
(1994).
Data were available on the reproductive and developmental effects of toxaphene, as well as
on growth or survival. In addition, the data set contained studies which were conducted over
chronic and subchronic durations. Laboratory experiments of similar types were not
conducted on a range of avian species and as such, inter-species differences among wildlife
species were not identifiable. There were no other values in the data set which were lower
than the benchmark value. Based on the avian data set for toxaphene, the benchmarks
developed from the Bush et al. (1977) .study were categorized as adequate.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) of 1.3 E-05 mg/1 (U.S. EPA,
1980) was selected as the benchmark protective of fish and aquatic invertebrates in the
generic freshwater ecosystem. It should be noted that a Final Residue Value (FRV) of 2.0 E-
7 mg/L was identified (U.S. EPA, 1986), however, it was not considered appropriate for a
benchmark value because residues and bioaccumulation are already taken into account by the
Thomann et al. (1992) model. Because the benchmark was based on a FCV derived for the
AWQC, this .benchmark is categorized as adequate.
August 1995
-------
APPENDIX B Toxaphene-5
Aquatic Plants: The toxicological benchmarks for aquatic plants were either: (l)ano
observed effects concentration (NOEC) or a lowest observed effects concentration (LOEC) for
vascular aquatic plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species
of freshwater algae, frequently a species of green algae (e.g., Selenastrum capricornutum).
Aquatic plant data was not identified for toxaphene and, therefore, no benchmark was
developed.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^.) to determine a protective chemical concentration
(Stephan, 1993). The EQP number is the chemical concentration that may be present in
sediment while still protecting the benthic community from the harmful effects of chemical
exposure. The FCV reported in the AWQC document for toxaphene (U.S. EPA, 1980) was
used to calculate a EQP number of 1.07 mg toxaphene /kg organic carbon. Assuming a mass
fraction of organic carbon for the sediment (f,,,.) of 0.05, the benchmark, for the benthic
community is 0.0535 mg/kg. Since the EQp number was based on a FCV established for the
AWQC, the sediment benchmark is categorized as adequate.
August 1995
-------
APPENDIX B
Toxaphene- 6
Table 1. Toxicologicai Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
RflpfS40nttiliv%
Specie*
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
Benchmark
Vtriutt'mg/kg-
day
1.24(p*)
0.69 (p*)
0.03 (a)
0.03 (a)
0.03 (a)
0.04 (a)
0.04 (a)
0.08 (a)
0.04 (a)
0.06 (a)
Study
Spmte*
rat
rat
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
Effect
dev
dev
dev
dev
dev
dev
dev
dev
dev
dev
Study V«hM
tngfl^diy
15
15
0.038
0.038
0.038
0.038
0.038
0.038
0.038
0.038
Owcrfptton
LOAEL
LOAEL
NOAEL
NCAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
SF
10
10
-
-
-
•
•
•
•
-
Original Sourc*
Chemoff and Carver,
1976
Chemoff and Carver,
1976
Bush et al.. 1977
Bush et al.,-1977
Bush et al., 1977
Bush et al., 1977
Bush et al., 1977
Bush et al., 1977
Bush et al., 1977
Bush et al., 1977
•Benchmark Category, a = adequate, p = provisional, i
of magnitude or more above the NEL or LEL for other
= interim; a '*' indicates that the benchmark value was an order
adverse effects.
August 1995
-------
APPENDIX B
Toxaphene • 7
Table 2. lexicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
nQprvocntativt
Spoctos
fish and aquatic
invertebrates
aquatic plants
benthic community
eencflflunt
V«to»
iftQnL
1.3E-05(a)
ID
5.35E-2 (a)
mg/kg sediment
Study
aquatic
organisms
aquatic
organisms
Qajjifhiilfui.
FCV
-
FCV x ^ .
Original Sourca
U.S. EPA, 1980
•
U.S. EPA, 1980
II.
'Benchmark Category, a = adequate, p = provisional, i = interim; a '*' indicates that the benchmark value was
an order of magnitude or more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem. .
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to toxaphene.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Chernoff and Carver, 1976) was used to derive the lexicological benchmark for
mammalian species representing the terrestrial ecosystem. The study value from Chernoff
and Carver (1976) study was divided by 10 to provide for a LOAEL-to-NOAEL safety factor.
This value was then scaled for species representative of a terrestrial ecosystem using a cross-
species scaling algorithm adapted from Opresko et al. (1994). Since Chernoff and Carver
(1976) documented reproductive effects from toxaphene exposure to female rats, female body
weights for each representative species were used in the scaling algorithm to obtain the
lexicological benchmarks. Based on the daia sei for toxaphene and because the benchmark is
August 1995
-------
APPENDIX B Toxaphene - 8
based on a LOAEL/10, the benchmarks developed from the Chernoff and Carver (1979) study
were categorized as provisional, with a "*" to indicate that adverse effects may occur at the
benchmark level..
Birds: No additional avian toxicity studies were identified for species representing the
terrestrial ecosystem. Thus, for the avian species representative of a terrestrial ecosystem., the
NOAEL of 0.038 mg/kg-day from the Bush et al. (1977) study was used as the benchmark
value. This value was then scaled for species representative of a terrestrial ecosystem using a
cross-species scaling algorithm adapted from Opresko et al. (1994);
Based on the avian data set for toxaphene, the benchmarks developed from Bush et al. (1977)
were categorized as adequate.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the LOEC values and then approximating the 10th
percentile. If there were 10 or fewer values for a chemical, the lowest LOEC was used. If
there were more than 10 values, the 10th percentiie LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation.
However, terrestrial plant studies were not identified for toxaphene and, as a result, a
benchmark could not be developed.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Toxaphene - 9
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
ftoplMMllfttlVV
SpMtas
deer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plant
soil community
flam m^m**mA
U9I1CMIIUUH
VrflM*
mgftg-ctay
3.1 (p')
3.2 (p-)
2-6 (p-)
1.1 (P*)
0.80 (p*)
0.77 (p-)
0.39 (p-)
0.04 (a)
0.06 (a)
0.06 (a)
0.07 (a)
0.05 (a)
ID
ID
Study
ScMfliiMr • :
w^^w^M
rat
rat
rat
rat
rat
rat
rat
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
white leghorn
chicken
-
•
•" •• rtttnt
.CROvE
dev
dev
dev
dev
dev
dev
dev
dev
dev
dev
dev
dev
-
-
Study
Vita*
m0r*r
d»y
15
15
15
15 '
15
15
15 .
0.038
0.038
0.038
0.038
.0.038
'
-
Doscfipfloii
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
/
LOAEL
LOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
•
.
SF
10
10
10
10
10
10
10
-
-
-
-
•.
-
-
Original Sourc«
Chemoff and Carver,
1976
Chemott and Carver,
1976
Chemoff and Carver,
1976
Chemoff and Carver.
1976
Chemoff and Carver,
1976
Chemoff and Carver,
1976
Chemoff and Carver,
"1976 '
Bush et al., 1977
Bush et al., 1977
Bush et al., 1977
Bush et al., 1977
Bushetal,, 1977
•
-
"Benchmark Category, a = adequate, p = provisional, i = interim; a '" indicates that the benchmark value was an order of magnitude or
more above the NEL or LEL for other adverse effects.
ID = Insufficient Data
August 1995
-------
APPENDIX B Toxaphene - 10
HI. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: trophic level 3
and 4 fish in the limnetic and littoral ecosystems, general fish (BCF only), aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. Each
' value is identified as whole-body or lipid-based and, for the generic aquatic ecosystems, the
biological uptake factors are designated with a "d" if the value reflects dissolved water
concentrations, and a "t" if the value reflects total surface water concentrations. For organic
chemicals with log K,,w values below 4, bioconcentration factors (BCFs) in fish were always
assumed to refer to dissolved water concentrations (i.e., dissolved water concentration equals
total water concentration). For organic chemicals with log K^ values above 4, the BCFs
were assumed to refer to total water concentrations unless the BCFs were calculated using
models based on the relationship between dissolved water concentrations and concentrations
in fish. The following discussion describes the rationale for selecting the biological uptake
factors and provides the context, for interpreting the biological uptake values presented in
Table 4.
i
As stated in section 5.3.2, the BAF/s for constituents of concern were generally estimated
using Thomann (1989) for the limnetic ecosystem and Thomann et al. (1992) for the littoral
ecosystem. Although these models were considered appropriate to estimate BAF/s for
toxaphene, comparison of the predicted BAF/s with the measured BAF/s suggested that the
predicted BAF/s may greatly underestimate the bioaccumulation potential of toxaphene. The
geometric mean BAF,d of two measured values on lake trout (32,388,515) was more than two
orders of magnitude above the predicted BAF,d (155,408) for trophic level 4 fish. The
discrepancy in measured vs. predicted BAF,ds is likely connected to the composition and
variability of toxaphene. Toxaphene is a mixture of. more than 175-179 components
(approximate molecular formula C10H,0Clg) produced by the chlorinatiori of camphene and, as
such, toxaphene batches may vary in their physicochemical properties. For example, log K,,w
values for toxaphene have been reported in shake-flask studies from 3.23 to 5.5 and the log
Kow calculated with the mechanistic SPARC model was 5.56 (Karickhoff, unpublished delta).
In addition, the high persistence of toxaphene may result in higher than expected
bioaccumulation in long-lived fish species and the heterogenous nature of the mixture may
impede clearance mechanisms (e.g., metabolism, excretion) in fish. Given the large
difference in measured and predicted values and the complex nature of toxaphene mixtures,
August 1995
-------
APPENDIX B Toxaphene-11
the geometric mean measured BAF/s was used for trophic level 4 fish in the limnetic
ecosystem. The trophic level 3 BAF/ was estimated by dividing the predicted ratio of BAF/'s
for trophic levels 4 and 3 (RBAF 4/3) into the measured BAF,d for trophic level 4 fish.
However, the higher than predicted bioaccumulation factor is probably related to the
persistence of toxaphene and, therefore, longer-lived fish would tend to accumulate more of
the chemical. Consequently, the RBAF 4/3 may overestimate the bioaccumulation potential
in smaller, shorter-lived fish in trophic level 3. The same BAF,ds that were used for trophic
levels 3 and 4 in the limnetic ecosystem were also used for the littoral ecosystem. Although
toxaphene is also highly persistent in sediments, these values should be interpreted with
caution since they may not adequately represent food web transfer in a sediment-based
ecosystem. The bioconcentration factor for fish was estimated as the geometric mean of 7
measured BCF,1 values presented in Stephan (1993). As with the BAF/, the predicted BCF/
was significantly lower (approximately an order of magnitude) than the geometric mean of
measured values. Therefore, the geometric mean of measured values was used as the BCF/
for fish. , -
The bioaccumulation factor for terrestrial vertebrates was the geometric mean of measured
values cited in Garten and Trabalka (1983). For terrestrial invertebrates and earthworms, the
bioconcentration factor was estimated as described in Section 5.3.5.2.3. Briefly, the
extrapolation method is applied to hydrophobic organic chemicals assuming that the
partitioning to tissue is dominated by lipids. Further, the method assumes that the BAFs and
BCFs for terrestrial wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of
Risks to Terrestrial Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993)
are of sufficient quality to serve as the standard. The beef biotransfer factor (BBFs) for a
chemical lacking measured data is compared to the. BBF for TCDD and that ratio (i.e.,
toxaphene BBF/TCDD BBF) is multiplied by the TCDD standard for terrestrial vertebrates, •
invertebrates, and earthworms, respectively. For hydrophobic organic constituents, the
bioconcentration factor for plants was estimated as described in Section 6.6.1 for above
ground leafy vegetables and forage grasses. The BCF is based on route-to-leaf translocation,
direct deposition on leaves and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
Toxapherie -12
Table 4. Biological Uptake Properties
ecological
receptor
limnetic trophic
level 4 fish
limnetic trophic
level 3 fish
fish
littoral trophic
level 4 fish
littoral trophic
level 3 fish
trophic level 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BAF
BAF
BCF
BAF
BAF
BAF
BAF
BCF
BCF
BCF
llpid-baaed or
whole-body
lipid
lipid
lipid
lipid
lipid
--
whole-body
whole-body
whole-body
whole-plant
value
42,349,000 (t)
40,988,942 (t)
407,300 (t)
42,349,000 (t)
40,988,942 (t)
-
0.37
0.0012
0.0094
0.05
•ource
based on geometric mean of
measured values from Swain et
al., 1986 as cited in Stephan,
1993
extrapolated from trophic level 4
BAF using RBAF 4/3 predicted
by Thomann, 1989
geometric mean of 7. measured
values presented in Stephan,
1993
same value assumed as in the
limnetic ecosystem
same value assumed as in the
limnetic ecosystem
insufficient data
geometric mean of values in
Garten and Trabalka. 1983
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDD
U.S. EPA, 1990e
d = refers to dissolved surface water concentration
t = refers to total surface water concentration
August 1995
-------
APPENDIX B Toxaphene-13
References
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Allen, A.L., L.D. Koller, and G.A. Pollock. 1983. Effect of toxaphene exposure on immune
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Bush, P.B., J.T. Kiker, R.K. Page, N.H. Booth, and O.J. Fletcher. 1977. Effects of graded
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APPENDIX B Toxaphene-14
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APPENDIX B Toxaphene-15
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August 1995
-------
APPENDIX B Toxaphene -16
Mayer, F.L., Jr., P.M. Mehrle, Jr., and W.P. Dwyer. 1975. Toxaphene effects on
reproduction growth, and mortaility of brook trout. Environ. Res. Lab., U.S.
Environmental Protection Agency, Duluth, MN. EPA-600/3-75-13. As cited in U.S. EPA
(U.S. Environmental Protection Agency). 1986. Ambient Water Quality Criteria for
Toxaphene - 1986. EPA 440/5-86-006. Criteria and Standards Division, Office of. Water
Regulations and Standards, Washington, D.C.
Mayer, F.L., P.M. Mehrle, and W.P. Dwyer. 1977. Toxaphene: Chronic Toxicity to Fathead
Minnows and Channel Catfish. Environ. Res. Lab., U.S. Environmental Protection
Agency, Duluth, MN. EPA-600/3-77-069. As cited in U.S. EPA (U.S. Environmental
Protection Agency). 1986. Ambient Water Quality Criteria for Toxaphene - 1986.
EPA 440/5-86-006. Criteria and Standards Division, Office of Water Regulations and
Standards, Washington, D.C.
Mehrle, P.M., and F.L. Mayer. 1975. Toxaphene effects on growth and development of
brook trout (Salvelinus fontinalis). J. Fish Res. Board Can. 32(5):609-613. As cited in
AQUIRE (AQUatic Toxicity_/nformation /?£trieval Database), Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Mehrle, P.M., M.T. Finley, J.L. Ludke, F.L. Mayer, and T.E. Kainse. 1979. Bone .
development in black ducks as affected by dietary toxaphene. Pestic. Biochem. Physiol.
10:168-173.
Nagy, K.A. 1987. Feild metabolism rate and food requirement scaling in mammals and birds.
Ecol. Mono. 57:111-128.
NCI (National Cancer Institute). 1977. Bioassay of toxaphene for possible carcinogenicity,
DHEW/PUB/NIH-79-837. Carcinogenesis Testing Program, Division of Cancer Cause
and Prevention, National Cancer Institute, Bethesda, MD.
Nebeler, A.V., W.L. Griffis, C.M. Wise, E. Hopkins, and J.A. Barbitta. 1989. Survival,
reproduction and bioconcentration in invertebrates and fish exposed to hexachlorobenzene.
Environmental Toxicology and Chemistry. 8:601-611.
Oliver and Nilmi. 1983. Bioconcentration of chlorobenzenes from water by rainbow trout:
correlations with partition coefficients and environmental residues. Environ. Sci. Technoi
17(5):287-291.
August 1995
-------
APPENDIX B Toxaphene -17
Oliver, E.G. 1987. Biouptake of chlorinated hydrocarbons from laboratory-spiked and field
sediments by oligochaete worms. Environ. Sci. Technol. 21:785-790.
Olson, K.L., F. Matsumura, and G.M. Boush. 1980. Behavioral effects on juvenile rats from
perinatal exposure to low levels of toxaphene and its toxic components, toxicant A, and
toxicant B. Arch. Environ. Contam. Toxicol. 9(2):247-257.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. lexicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. .U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Peakall, D.B. 1976. Effects of toxaphene on hepatic enzyme induction and circulating
steroid levels in the rat. Environ. Health Perspect. 13:117-120.
Pollock, G.A, and W.W. Kilgore. 1978. Toxaphene. Residue Rev. 69:87-140.
RTECS (Registry of Toxic Effects of Chemical Substances). 1994. National Institute for
Occupational Safety and Health. Washington, DC.
Richardson , M.E., M.R. Spivey Fox, and B.E. Fry. 1974. Pathological changes produced in
Japanese quail by ingestion of cadmium. J. Nutr. 104:323-338.
Ros'eberry and,Klimistra. 1971. Annual weight cycles in male and female bobwhite quail. Auk
88:116-123.
Sanders, H.O. 1980. Sublethal Effects of Toxaphene on Daphnids, Scuds, and Midges.
U.S. Environmental Protection Agency Report 600/3-80-006. As cited in Eisler, R. and J.
Jacknow. 1985. Toxaphene hazards to fish, wildlife, and invertebrates: a
synoptic review. U.S. Fish Wildlife Servic. Biol. Rep. 85 (1.4). pp.26.
Schimmel, S.C., J.M. Patrick, and J. Forester. 1977. Uptake and toxicity of toxaphene in
several estuarine organisms. Arch. Environ. Contam. Toxicol. 5:353-367. As cited in
Stephan. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative, PB93-154672, Environmental Research Laboratory,
Office of Research and Development, Duluth, MN.
August 1995
-------
APPENDIX B Toxaphene -18
Schwetz, B.A., J.M. Morris, R.J. Kociba, P.A..Keeler, R.F. Cornier, and P.J. Gehring. 1974.
Reproduction study in Japanese quail fed hexachlorobenzene for 90 days. Toxicol. appi.
Pharmacol. 30:255-265.
* • •
Stephan, C.E. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
Laboratory, Office of Research and Development, Duluth, MN.
Suter II, G.W. and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, D.C.
Swain, W.R., M.D. Mullin, and J.C. Filkins. 1986. Long range transport of toxic organic
contaminants to the North American Great Lakes. In: Ryans, R.C., ed. Problems of
Aquatic Toxicology, Biotesting and Water Quality Management. EPA-600/9-86-024.
National Technical Information Service, Springfield, VA. pp. 107-121. As cited in
Stephan. 1993. Derivations of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative, PB93-154672, Environmental Research Laboratory,
Office of Research and Development, Duluth, MN.
Thomann, R.V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
food chains. Environ. Sci. Technol. 23(6):699-707.
Thomann, R.V., J.P. Connolly, and T.F. Parkerton. 1992. An equilibrium model of organic
chemical accumulation in aquatic food webs with sediment interaction. Environmental
Toxicology anaI Chemistry 11:615-629. -
U.S. EPA (U.S. Environmental Protection Agency). 1976. Criteria Document for Toxaphene.
EPA/440/9-76/014. Washington, D.C.
U.S. Department of Health, Education, and Welfare. 1979. Bioassay of Toxaphene for
Possible Carcinogenicity. Carcinogenesis Testing Program, Division of Cancer Cause
and Prevention, Nationals Cancer Institute. DHEW Publication No. (NIH) 79-837.
/
U.S. EPA (Environmental Protection Agency). 1980. Ambient Water Quality Criteria for
Chlorinated Benzenes. EPA-440/5-80-028. Criteria and Standards Division, Washington,
D.C.
August 1995
-------
APPENDIX B Toxaphene - 19
U.S. EPA (U.S. Environmental Protection Agency): 1980. Ambient Water Quality Criteria
for Toxaphene. EPA 440/5-80-076. Criteria and Standards Division, Office of Water
Regulations and Standards, Washington, D.C.
U.S. EPA (Environmental Protection Agency. 1984. Health Effects Assessment for
Hexachlorobenzene. Environmental Criteria and Assessment Office, Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1986. Ambient Water Quality Criteria
for Toxaphene - 1986. EPA 440/5-86-006. Criteria and Standards Division, Office of
Water Regulations and Standards, Washington, D.C. v
U.S. EPA (U.S. Environmental Protection Agency). 1987. Health Effects Assessment for
Toxaphene. EPA/600/8-88/056. Environmental Criteria and Assessment Office, Office of
Health and Environmental Assessment, Cincinnati, OH.
U.S.EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. EPA/600/6-87/008.
Environmental Criteria and Assessment Office, Office of Health and Environmental
Assessment, Office of Research and Development, Cincinnati, OH.
U.S. EPA (U.S. Environmental Protection Agency). 1989. Ambient Water Quality Criteria
Document addendum for Toxaphene. (Draft Rep. (Final).). PB91 - 161588. Criteria and
Assessment Office. Washington, D.C.
U.S. EPA (U.S. Environmental Protection Agency). 1990e. Methodology for Assessing
Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final.
Office of Health and Environmental Assessment, Washington, D.C. January.
U.S. EPA (U.S. Environmental Protection Agency). 1993b. Wildlife Criteria Portions of the
Proposed Water Quality Guidance for the Great Lakes System. EPA-822-R-93-006.
Office of Science and Technology, Office of Water, Washington, D.C.
U.S. EPA (Environmental Protection Agency). 1993c. Technical Basis for Deriving Sediment
Quality Criteria for Nonionic Organic Contaminants for the Protection of Benthic
Organisms by Using Equilibrium Partitioning. EPA/822-R-93/011. Office of Water, "
Washington, D.C.
August 1995
-------
APPENDIX B Toxaphene-20
Vonrumher, R., E.W. Lawless, and A.F. Meiners. 1974. Production, distribution, use, and
environmental impact potential of selected pesticides. U.S. Environmental Protection
Agency, PB-238-795. As cited in Pollock, G.A, and W.W. Kilgore. 1978. Toxaphene.
Residue Rev. 69:87-140.
Vos, J.G., H.L. Van Der Maas, A. Musch and E. Ram. 1971. Toxicity of
Hexachlorobenzene in Japanese Quail with Special Reference to Porphyria, Liver Damage,
Reproduction, and Tissue Residues. Toxicology and Applied Pharmacology, 18:944-957.
August 1995
-------
Terrestrial Biological Uptake Measures - Toxaphene
CAS No. 8001-35-2
Chemical .
Name
toxaphene
toxaphene
toxaphene
toxaphene
Species
sheep
poultry
cow
plants
B-factor
(BCF. BAF.
BMF)
"%,.
BAF
BAF
BAF
BCF
Value
0.046
10.72
0.1
11
Measured
or
predicted
(m.p)
p
P
P
P
Units
kg fat/kg diet
kg fat/kg diet
kg fat/kg diet
(ug/g DW
plant)/(ug/g soil)
Reference
Garten and Tralbalka,
1983 .
Garten and Tralbalka,
1983
Garten and Tralbalka,
1983
U.S. EPA, 1990e
Comments
-------
Freshwater Biological U^ xe Measures - Toxaphene
Cas No. 8001-35-2
Chemical
Name
toxaphene
toxaphene .
toxaphene
toxaphene
toxaphene
Species
[fish
fish
lake trout
lake trout
fish
B-factor
(BCF. BAF.
BMF)
BCF
BCF
BAF
BAF
BCF
Value
22107
4103
839,695
213,582
467
Measured or
predicted
(m.p)
m
m
m
m
P
Units
NS
NS
NS
NS
NS
Reference
Goodman, 1986 as cited in
Stephan, 1993
Goodman, 1986 as cited in
Stephan, 1993
Swain et ai. 1986 as cited
in Stephan, 1993
Swain etal, 1986 as cited
in Stephan, 1993
Stephan, 1993
Comments
Normalized to 1 .0% lipid; adults.
Normalized to 1 .0% lipid; juveniles.
Normalized to 1 .0% lipid.
Normalized to 1 .0% lipid.
Normalized to 1.0% lipid.
NS = Not specified
-------
Freshwater Biological Uptake Measures - Toxaphene
Cas No. 8001-35-2
Chemical .
Name
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
Species
channel
cattish
channel
catfish
channel
catfish
channel
catfish
channel
catfish
channel
catfish
channel
catfish
•
fish
fish
fish
tish
B-factor
(BCF. BAF,
BMR
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF
BCF
BCF
BCF
Value
6,111
1,477
12,000
8,889
2,535
8,298
2,895
22949
2633
5578
15235
Measured or
predicted
(m,p)
m
m
m
m
m
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Mayer et a!., 1977 as cited
in U.S. EPA, 1986
Mayer et al., 1977 as cited
in U.S. EPA, 1986
Mayer et al., 1977 as cited
in U.S. EPA, 1986
Mayer et al., 1977 as cited
in U.S. EPA, 1986
Mayer et al., 1977 as cited
in U.S. EPA, 1986
Mayer et al., 1977 as cited
in U.S. EPA, 1986
Mayer et at., 1977 as cited
in U.S. EPA, 1986
Schimmel et al., 1977 as
cited in Stephan, 1993 '
Goodman et al., 1978 as
cited in Stephan, 1993
Goodman, 1986 as cited in
Stephan, 1993
Goodman, 1986 as cited in
Stephan. 1993
Comments
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) = 30;
BCFs and BAFs divided by 1 .8%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) = 30;
BCFs and BAFs divided by 8.8%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) = 50;
BCFs and BAFs divided by 8.2%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) = 60;
BCFs and BAFs divided by 2.7%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) =75;
BCFs and BAFs divided by 71%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) =90;
BCFs and BAFs divided by 4.7%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.049-0.630; Duration (days) =100;
BCFs and BAFs divided by 7.6%
lipids; normalized to 1% lipid.
Normalized to 1 .0% lipid; juveniles.
Normalized to 1 .0% lipid; juveniles.
Normalized to 1 .0% lipid; juveniles.
Normalized to 1 .0% lipid; adults.
-------
Freshwater Biological Uj. .e Measures - Toxaphene
Cas No. 8001-35-2
Chemical .
Name
oxaphene
toxaphene
loxaphene
toxaphene
toxaphene
toxaphene
oxaphene
toxaphene
loxaphene
toxaphene
toxaphene
toxaphene
Species
brook trout
brook trout
brook trout
Drook trout
brook trout
jrook trout
(athead
minnow
arook trout
fathead
minnow
fathead
minnow
fathead
minnow
fathead
minnow
B-factor
(BCF, BAF,
BMP)
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
BCF
BCF or BAF
BCF or BAF
BCF or BAF
BCF or BAF
Value
12,000
4,200
18,000
9,400
6,400
3,100
90,000
71,500
3,077
3,860
5,484
2,926
Measured or
predicted
(m.p)
m
m
m
m
m
m
m
m
m
m
m
m
Units
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
NS
Reference
Mayer et a!., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mayer et al!, 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mehrfe and Mayer, 1 975 as
cited in AQUIRE, 1995
Mayer et at., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1975 as cited
in U.S. EPA, 1986
Mayer et al., 1977 as cited
in U.S. EPA. 1986
Comments
Measured cone, in water (ug/L) =
0.039-0.139; Duration (days) = 60.
Measured cone, in water (ug/L) =
0.039-0.139; Duration (days) = 60.
Measured cone, in water (ug/L) =
0.039-0.139; Duration (days) = 90.
Measured cone, in water (ug/L) =
0.039-0.139; Duration (days) = 140.
Measured cone, in water (ug/L) =
0.039-0.139; Duration (days) =161.
Measured cone, in water (ug/L) =
0.039-0.139; Duration (days) = 161.
Measured cone, in water (ug/L) =
0.055-0.621; Duration (days) = 150.
Fry; 15 day test.
Measured cone, in water (ug/L) =
0.013-0.173; Duration (days) = 30;
BCFs and BAFs divided by 5.2%
lipids = normalized to 1 % lipid .
Measured cone, in water (ug/L) =
0.013-0.173; Duration (days) = 30;
BCFs and BAFs divided by 5.7%
lipids; normalized to 1% lipid.
Measured cone, in water (ug/L) =
0.013-0.173; Duration (days) = 30;
BCFs and BAFs divided by 9.3%
lipids; normalized to 1 % lipid.
Measured cone, in water (ug/L) =
0.013-0.173; Duration (days) = 295;
BCFs and BAFs divided by 2.7%
lipids; normalized to 1% lipid.
-------
Freshwater Toxicity - Toxaphene
Cas No. 8001-35-2
Chemical
Name
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
Species
aquatic
organisms
aquatic
organisms
daphnid
amphipod
midge, larva
fathead
minnow
channel
catfish
brook trout
fathead
minnow
channel
catfish
bluegill
rainbow trout
Endpolnt
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
Description
FCV
AWQC
CV
CV
CV
CV
CV
CV
LC50
LC50
LC50
LC50
Value
0.013
0.0002
0.07-0.12
0.13-0.25
1.0-3.2
0.025-0.054
0.129-0.299
<0.039
5.0 - 23
(10.3)
0.8-16.5
(4.58)
2.4-4.7
(3.33)
8.4
Units
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
ug/L
Test type
(static/ flow
through)
NA
NA
NS
NS
NS
NS
partial life
cycle test
partial life
cycle test
NS
NS
NS
NS
Exposure
Duration/
Timing
NS
NS
NS
NS
NS
NS
NS
NS
16 days
4 days
4 days
4 days
Reference
U.S. EPA, 1980
U.S. EPA, 1986
Sanders, 1980 as cited in
Eisler, 1985
Sanders, 1980 as cited in
Eisler, 1985
Sanders, 1980 as cited in
Eisler, 1985
Mayer etal., 1977
Mayer etal., 1977
Mayer, etal., 1975
AQUIRE, 1995
AQUIRE, 1995
AQUIRE, 1995
AQUIRE. 1995 -
Comments
Based on the Final
Residue Value; No FCV
reported
Arthropod
Arthropod
Arthropod
Critical life stage end
points: embryo, larval, and
early juvenile; growth.
Critical life stage end
points: embryo, larval, and
early juvenile; growth.
Critical life stage end
points: embryo, larval, and
early juvenile; growth.
recalculated value
NS = Not specified
-------
Terrestrial Tox.. .ty - Toxaphene
Cas No. 8001-35-2
Chemical
Name
toxaphene
loxaphene
toxaphene
toxaphene
toxaphene
toxaphene
Species
pheasant
gray partridge
sandhill crane
homed lark
mule deer
domestic goat
Endpolnt
mort.
mort.
mort.
mort.
mort.
mort.
Benchmark
(NOAEL,
NOEL.
LOAEL,
LOEL, PEL,
LD50)
LD50
LD50
LD50
LD50
L050
LD50
Value
40
23.7
100-316
581
139 - 240
>160
Units
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral,
B.C., I.V., l.p.,
injection)
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
NS
NS
NS
NS
NS
NS
Reference
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Comments
•
NS = Not specified
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid - 1
Toxicologies! Profile for Selected Ecological Receptors
2,4,5-Trichlorophenoxyacetic acid
CasNo.: (93-76-5)
Summary: This profile on 2,4,5-Trichlorophenoxyacetic acid (2,4,5-T) summarizes the
lexicological benchmarks and biological uptake measures (i.e., bioconcentration,
bioaccumulation, and biomagnification factors) for birds, mammals, daphnids and fish, aquatic
plants and benthic organisms representing the generic freshwater ecosystem and birds,
mammals, plants, and soil invertebrates in the generic terrestrial ecosystem. Toxicological
benchmarks for birds and mammals were derived for developmental, reproductive or other
effects reasonably assumed to impact population sustainability. Benchmarks for daphnids,
benthic organisms, and fish were generally adopted from existing regulatory benchmarks (i.e.,
Ambient Water Quality Criteria). Bioconcentration factors (BCFs), bioaccumulation factors
(BAFs) and, if available, biomagnification factors (BMFs) are also summarized for the
ecological receptors, although some BAFs for the freshwater ecosystem were calculated for
organic constituents with log K^ between 4 and 6.5. For the terrestrial ecosystem, these
biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at
the end of this profile. This profile represents the most current information and may differ
from data presented in the technical support document for the Hazardous Waste
Indentification Rule (HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^ for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
•*
Study Selection and Calculation of Toxicological Benchmarks
Mammals: No suitable subchronic or chronic studies were found which reported dose-
response data for mammalian wildlife. However, lexicological studies involving 2,4,5-T
exposure to mammals have been conducted using laboratory animals. Two studies were
identified for consideration of benchmarks. Smith et al. (1981) conducted a three-generation
chronic study on rats exposed to 2,4,5-T. In this study, Sprague-Dawley male and female rats
August 1995
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid - 2
were administered 3, 10, or 30 mg/kg-day dietary 2,4,5-T for 90 days prior to mating. A
consistant tendency toward a reduction in neonatal survival was found at the dose level of 30
mg/kg/day. No other effects on reproductive capacity were seen. Based on these results, a
LOAEL of 30 mg/kg-day and a NOAEL of 10 mg/kg-day were inferred. Collins and
Williams (1971) examined reproductive and teratogenic effects by administering 2,4,5-T to
pregnant female golden hamsters via oral intubation on days 6 through 10 of gestation. The
hamsters were dosed with 40, 80, and 100 mg/kg-day with a significant dose-related decrease
in fetal viability. There was also increased levels of embryonic mortality and the number of
live born with hemorrhages. There were also no malformations produced below the 100
mg/kg-day level. This resulted in an inferred LOAEL of 40 mg/kg-day for the reproductive
effects.
The LOAEL in the Collins and Williams (1971) study was chosen to derive the lexicological
benchmark because (1) chronic exposures were administered via oral intubation, (2) it focused
on reproductive toxicity as. a critical endpoint, and (3) the study contained dose-response
information. The study by Smith et al. (1981) was not selected for the derivation of a
benchmark because of experimental conditions that may have confounded the results. For
example, during gestation of the F3b liters, some of the adult females in the control and
treated groups suffered water deprivation at various periods as a result of malfunctioning of
the automatic watering system. There were also incidences of accidental death of treatment
animals and the controls in the study were not raised in consistant conditions throughout the
generations of the study. Therefore, the infer-d LOAEL of 40 mg/kg-day from the Collins
and Williams (1971) study was used for the derivation of the mammalian benchmarks.
The study value from Collins and Williams (1971) was divided by 10 to provide a LOAEL-
to-NOAEL safety factor. This value was then scaled for species representative of a
freshwater ecosystem using a cross-species scaling algorithm adapted from Opresko et al.
(1994): •
bw
Benchmark. = NOAEL, x _ 1
where NOAEL, is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is thtf^feody weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Collins and Williams (1971) study documented reproductive effects from 2,4,5-T exposure to
August 1995
-------
Terrestrial Toxicity - Toxaphene
Cas No. 8001-35-2
Chemical
Name
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
Species
male rat
female rat
female rat
CD-1 mice
4 male and 14
female Swiss
white mice
monkey
N
domestic white
leghorn
chickens
ducklings
Endpolnt
rep
rep
rep
rep
rep
chronic
dvp, rep
dvp
Benchmark
(NOAEL,
NOEL,
LOAEL,
LOEL, PEL,
LDSO)
NOAEL
NOAEL
LOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
Value
37
49
27
35
25
0.7
0.038
44
-
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
ppm
ppm
mg/ka-day
mq/kq-day
Exposure
Route (oral,
8.C., I.V., l.p.,
Injection)
oral (feed)
oral, (feed)
oral (feed)
oral (qavaqe)
oral
oral
oral
oral
Exposure
Duration/Timing
48wks, ad lib
48wks, ad lib
80wks, ad lib
Days 7- 16 of
gestation
5 to 6 generation
study
2-year period
30 weeks
90 days
Reference
Chuetal., 1988
Chuetal., 1988
NCI, 1977
Chemoff and Carver,
1976
Keplingeret a!., 1970
Vonrumher et al., 1974
as cited in Pollock and
Kilgore, 1978
Bushetal., 1977
Mehrteetal., 1979
Comments
No value sited as having an
impact on rep, so NOAEL could
actually be higher (implication tr'ia
there are no repro effects
attributed to toxaphene).
No value sited as having an
impact on rep., so NOAEL could
actually be higher (implication tha
there are no repro effects
attributed to toxaphene).
Vaginal bleeding. No dose
response data given in the
summary.
This dose 'did not significantly
alter egg production, hale! lability,
or fertility, although some bone
deformation and kidney lesions
were recorded in adults." Doses
were 0, 0.5, 5, 50, and 100 ppm.
Diets contained 10 or 50 mo/kg.
-------
i erresinai i o, y - i oxapnene
Cas No. 8001-35-2
Chemical
Nam*
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
SMCtM
8 male and 1 6
female SD rats
16-39 CD
pregnant rats
5 pregnant CD
rats
rats
rat
rat
rat
Endpolnt
rep
rep
dev
behv
rep
behav
immuno. dev
Benchmark
(NCAEL,
NOEL,
LOAEL,
LOEL, PEL,
LD50)
NOAEL
LOAEL
LOAEL
LOAEL
NOAEL
NOAEL
NOAEL
Value
7.3
15
12.5
0.05
12Q
6
1.9
Units
mg/kg-day
mg/kg-dav
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-dav
mq/kq-dav
Exposure
Route (oral,
S.C., I.V., l.p.,
Inlectlon)
oral
oral (gavage)
oral (gavage)
oral
oral
(capsule)
oral (gavage)
oral (feed)
Exposure
Duration/Timing
3 generation
study
Days 7- 16 of
gestation
Days 7- 16 of
gestation
Day 5 of
gestation to 3
months
postpartum
once
15 days; GD 7-21
9.5wks (before
breeding, during
preg and during
lactation)
Reference
Kennedy etal., 1973
Chemoff and Carver,
1976
Kavlocketal., 1982
_
Olson etal., 1980
Peakall, 1976
Crowder et al., 1980
Allen etal., 1983
Comments
No dose-response data. The
animals were started on the diet
at 28 days old. Doses were 25
and 100 ppm 100 ppm dose
converted using body wt. of 0.458
kg and 0.033 kg food/day
(U.S.EPA, 1988)
Inferior swimming ability.
No dose response information
provided in the summary.
Behavioral changes in pups
(impaired righting reflex). Only
single dose tested.
Immunosupression in offspring
(reduced phagocytic ability in
macrophages, deer humoral
antibody response). Questionable
correlation with population effects
Dose-response unclear (0, 10,
100,200).
-------
Terrestrial Toxicity - Toxaphene
Cas No. 8001-35-2
Chemical
Name
toxaphene
toxaphene
toxaphena
toxaphene
toxaphene
loxaphe'ne •
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
loxaphene
toxaphene
Species
Northern
bobwhite, 3-
day old
rat
mouse
dog
rabbit
guinea pig
hamster
duck
fulvous
whistling duck
mallard
duckling
mallard
sharp tailed
grouse
bobwhite
California quail
Endpolnt
behv
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort.
mort. >
mort.
mort.
mort.
Benchmark
(NOAEL,
NOEL,
LOAEL,
LOEL, PEL,
LO50)
LOAEL
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
LD50
Value
10
50
112
15
75
250
200
31
99
30.8
70.7
19.9
85.5
23.7
Units
mg/kg-diet
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
mg/kg-body
wt.
Exposure
Route (oral,
8.C., I.V., l.p.,
Injection)
oral
oral
oral
oral
oral
oral
oral .
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration/Timing
20 weeks
NS
NS
NS
NS
NS
NS
NS
NS x .
NS
NS
NS
NS
NS
Reference
KreiUer, 1980
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
U.S. EPA, 1993b
Comments
50% more behavonal errors than
controls on initial testing to
stimulus pattems(30 days after
exposure) In the second test,
there was no difference between
experimentals and controls.
Doses given were 10 or 50
mg/kg.
t
-
-------
Terrestrial To* .y - Toxaphene
Cas No. 8001-35-2
Chemical
Name
toxaphene
toxaphene
toxaphene
toxaphene
toxaphene
Species
American black
ducks
American black
ducks
Mallard eggs
Ring-necked
pheasants
Northern
bobwhite
Endpolnt
rep, dvp
aehv
emb, dvp
rep. dvp
dvp, behv
Benchmark
NOAEL,
NOEL,
LOAEL,
LOEL, PEL,
LD50)
NOAEL
NOAEL
NOAEL
LOAEL
AEL
Value
50
50
1.12
100
5
Units
mg/kg-diet
mg/kg-diet
kg/ha
mg/kg-diet
mg/kg-diet
Exposure
Route (oral,
Infection)
oral
oral
application
oral
oral
Exposure
Duration/Timing
1 9 months
(lasting two
breeding
seasons)
NS
NS
NS
4 months
Reference
Haseltine et al , 19BO
Heinz and Finley, 1 978
Hoffman and Eastin,
1982
Genelly and Rudd,
1958 as cited in U.S.
EPA, 1976
Pollock and Kilgore,
1978
Comments
No effects on survival, egg
production, fertility, hatchability,
eggshell thickness, or growth and
survival of young at 10 or 50 '
mg/kg. No dose-response.
'There was no change in
avoidance behavior of this
species, which, if interrupted, is
considered life-threatening."
Doses 0, 10, and 50 ppm. No
dose-response.
Studies have shown that if this
application rate is exceeded,
which is normally the case, then
severe embryotoxic effects,
including dislocated joints and
poor growth may occur.
Both dose levels, 100 and 300
ppm, caused greater mortality in
young pheasants during the first 2
weeks after hatching.
Effects: thyroid hypertrophy and
interference with the ability of
bobwhites to discriminate
patterns.
-------
APPENDIX 8 , 2,4,5-Trichlorophenoxyacetic acid - 3
pregnant female hamsters, the mean female body weights of representative species were used
in the scaling algorithm to obtain the toxicological benchmarks.
Data were available on the reproductive and developmental effects of 2,4,5-T, as well as
growth or chronic survival. In addition, the data set contained studies which were conducted
over chronic and subchronic durations and during sensitive life stages. Based on the data set
for 2,4,5-T and because the benchmark is based on a LOAEL/10, the benchmarks developed
from the Collins and Williams (1971) study were categorized as provisional.
Birds: No suitable studies were found for 2,4,5-T toxicity in avian species associated with the
freshwater ecosystem. ,
Fish and aquatic invertebrates: No AWQC or Final Chronic Value (FCV) was available for
2,4,5-T. Therefore, a Secondary Chronic Value (SCV) of l.OE-02 mg/L was calculated using
the Tier II methods described in Section 4.2.5. Because the benchmark was derived using the
the Tier II method, it was categorized as interim.
Aquatic plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g. duckweed) or (2) an effective concentration (EC,,) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). For 2,4,5-T
there was insufficient data for the development of a benchmark value.
Benthic community: Benchmarks for the protection of benthic organisms were determined
using the Equilibrium Partition (EQP) method. The EQP method uses a Final Chronic Value
(FCV) or other chronic water quality measure, along with the fraction of organic carbon and
the octanol-carbon partition coefficient (K^.) to determine a protective sediment concentration
(Stephan, 1993). The EQP number is the chemical concentration that may be present in
sediment while still protecting the benthic community from harmful effects from chemical
exposure. Because no FCV was available, a Secondary Chronic Value was calculated as
described in Section 4.3.5. The SCV reported for 2,4,5-T was used to calculate an EQP
number 1.24E+01mg 2,4,5-T per kg organic carbon. Assuming a mass fraction of organic
carbon for the sediment (f^.) of 0.05, the benchmark for the benthic community is 6.2E-01 mg
2,4,5-T per kg of sediment. Because the EQp number was set using a SCV derived using the
Tierll method , it was categorized as yjterim.
August 1995
-------
APPENDIX B
2,4,5-Trichlorbphenoxyacetic acid - 4
Table 1. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
ItnM^UM
mink
river otter
bald eagle
osprey
great blue heron
mallard
lesser scaup
spotted sandpiper
herring gull
kingfisher
*«*»*<
1 «*
2.76 (p)
1.54(p)
ID
10
10
10
10
10
10
ID
•. JSSjfc
^•CMB
l> / "•
*>••" -t.
hamster
hamster
-
-
•
•
-
-
-
-
C::MM*-'
H^mn*
?J? r :
fe **•<,
rep
rep
-
-
-
-
-
-
.
- -
-------
APPENDIX B
2,4,5-Trichlorophenoxyacetic acid - 5
Table 2. lexicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
fish and aquatic
invertebrates
aquatic plants
benthic community
1.0E-020)
ID
6.2E-01 (i)
aquatic
organisms
aquatic
organisms
SCV
SCVxK,,
AQUIRE
AOUIRE
II.
•Benchmark Category, a 3 adequate, p = provisional, I = interim; a "' indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
ID = insufficient data
Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C^) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, no suitable
subchronic or chronic studies were found for mammalian wildlife exposure to 2,4,5-T.
Because of the lack of additional mammalian toxicity studies, the same surrogate-species
study (Collins and Williams, 1971) was used to derive the 2,4,5-T lexicological benchmark
for mammalian species representing the terrestrial ecosystem. The study value from the
Collins and Williams (1971) study was divided by 10 to provide for a LOAEL-to-NOAEL
safety factor. This value was then scaled for species in the terrestrial ecosystem using a
cross-species scaling algorithm adapted from Opresko et al. (1994). Since the Collins and
Williams (1971) study documented rejrgpductive effects from 2,4,5-T exposure to pregnant
female hamsters, the mean female body-weights of each representative species were used in
the scaling algorithm to obtain the lexicological benchmarks.
August 1995
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid - 6
Based on the data set for 2,4,5-T and because the benchmark is based .on a LOAEL/10, the
benchmarks developed from the Collins and Williams (1971) study were categorized as
provisional, as in the aquatic ecosystem.
Birds: No suitable studies were found for 2,4,5-T toxicity in avian species associated with
the terrestrial ecosystem.
Plants: Adverse effects levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks
were selected by rank ordering the LOEG values and then approximating the 10th percentile.
If there were 10 or fewer values, the 10* percentile LOEC was used. Such LOECs applied to
reductions in plant growth, yield reductions, or other effects reasonably assumed to impair the
ability of a plant population to sustain itself, such as a reduction in seed elongation. \
However, studies were not identified for benchmark development for 2,4,5-T.
Soil Community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
j
-------
APPENDIX B
2,4,5-TricrVorophenoxyacetic acid - 7
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
Hfl^fBVttfHBDVv
ft^^fifa*-*
dear mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red-tailed hawk
American kestrel
Northern
bobwhrte
American robin
American
woodcock
plants .
soil community
DmaiuiMrfc
V»to«*
^__^^j^— ^
uiymuiy -
6.81 (p)
7.01 (p)
5.69 (p)
2.41 (p)
1.79(p)
1.72(p)
0.86 (p)
ID
- ID
ID
ID
ID
ID
ID
, Stud?
• SpaolM
hamster
hamster
hamster
hamster
hamster
hamster
hamster
-
-
-
-
-
, -
-
Htact
rep
rep
rep
rep
rep
rep
rep
•
-
• • -
-
-
. -
•
Study
; v*»
"flflpBs.
^,
40
40
40
40
40
40
40
-
-
-
-
-
-
-
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
LOAEL
.
,
- • '
-
-
-
*>.
10
10
10
10
10
10
10
-
•
-•.
• •
-
-
-
{M£|M§«MVW*
Collins and
Williams, 1971
Collins and
Williams, 1971
Collins and
Williams. 1971
Collins and
Williams, 1971
Collins and
Williams, 1971
Collins and
Williams, 1971
Collins and
Williams, 1971
•
-
-
-
-
-
'Benchmark Category, a = adequate, p = provisional, i - interim; a "' indicates that the benchmark value was an order
of magnitude or more above the NEL or L.%Jor other adverse effects.
ID = insufficient data
August 1995
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid - 8
III. Biological Uptake Measures
This section presents biological uptake measures (i.e. BCFs, BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for selected ecological receptor categories: aquatic
invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and plants. For the
generic aquatic ecosystems, the BCF value is identified as whole-body or lipid-based and
designated with a "d" if the value reflects dissolved water concentrations, and a "t" if the
value reflects total surface water concentrations. For organic chemicals with log K,,w values
below 4, bioconcentration factors (BCFs) in fish were always assumed to refer to dissolved
water concentrations (i.e., dissolved water concentration equals total water concentration).
The following discussion describes the rationale for selecting the biological uptake factors and
provides the context for interpreting the biological uptake values presented in Table 4.
The bioconcentration factor for fish was estimated from the Thomann (1989) model (i.e., log
KVV - dissolved BCF/) because: (1) no appropriate measured values were identified, (2) the
BCF was in close agreement with predicted BCFs based on other methods (i.e., regression
equations), and (3) there were no data (e.g., metabolism) to suggest that the log K^ = BCF,d
relationship deviates for 2,4,5-T (log K^ = 3.13). As stated in section 5.3.2, the dissolved
bioconcentration factor (BCF," ) for organic chemicals with log K^ below 4 was considered
to be equivalent to the total bioconcentration factor (BCF,1) and, therefore, adjusting the BCF,d
by the dissolved fraction (fd) was not necessary.
The bioaccumulation/bioconcentration factors for terrestrial vertebrates, invertebrates, and
earthworms were estimated as described in Section 5.3.5.2.3. Briefly, the extrapolation
method is applied to hydrophobic organic chemicals assuming that the partitioning to tissue is
dominated by lipids. Further, the method assumes that the BAFs and BCFs for terrestrial
wildlife developed for 2,3,7,8-TCDD in the Revision of Assessment of Risks to Terrestrial
Wildlife from TCDD and TCDF in Pulp and Paper Sludge (Abt, 1993) are of sufficient
quality to serve as the standard. The beef biotransfer factor (BBFs) for a chemical lacking
measured data is compared to the BBF for TCDD and that ratio (i.e., 2,4,5-T BBF/TCDD
BBF) is multiplied by the TCDD standard for terrestrial vertebrates, invertebrates, and
earthworms, respectively. For hydrophobic organic constituents, the bioconcentration factor
for plants was estimated as described'!* Section 6.6.1 for above ground leafy vegetables and
forage grasses. The BCF is based on route-to-leaf translocation, direct deposition on leaves
and grasses, and uptake into the plant through air diffusion.
August 1995
-------
APPENDIX B
2,4,5-Trichlorophenoxyacetic acid - 9
Table 4. Biological Uptake Properties
•ssr
fish
littoral
trophic level 2
invertebrates
. terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF, BAF, or
BSAF
BCF
-
BAF
BCF
BCF
BCF
ffa^^^^^^Mfi' t&"
wli^te&^riv*
lipid
-
whole-body
whole-body
whole-body
whole-plant
„ «*«• .
1 ,350 (d)
ID
1.7E-05
1.6E-05
1.3E-04X
6.0E-01
... • -"-
predicted; Thomann, 1989
•
estimated based on beef
biotransfer ratio with 2,3,7.8-
TCOO
estimated based on beef
biotransfer ratio with 2.3,7,8-
TCDD
estimated based on beef
biotransfer ratio with 2,3,7,8-
TCDO
U.S. EPA. 1990e
d - refers to dissolved surface water concentration
t = . refers to total surface water concentration
ID = insufficient data
August 1995
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid • 10
References
AQUIRE ( AQUatic Toxicity /nformation /?£trievaJ Database). 1995. Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Collins, T. F. X., C. H. Williams. 1971. Teratogenic studies with 2,4,5-T and 2,4-D in the
hamster. Bulletin of Environmental Contamination and Toxicology. 6:559-567.
Courtney, K. D., D. W. Gaylor, M. D. Hogan, and H. L. Falk. 1970. Teratogenic evaluaution
of 2,4,5-T. Science. 168:864-866.
Crampton, M.A., and L.J. Rogers. 1982. Low doses of 2,4,5-trichlorophenoxyacetic acid are
behaviorally teratogenic to rats. Experienta 39:891-892.
Dougherty, W.J., F. Coulstbn, L. Golberg. 1976. The Evaluation of the Teratogenic Effects
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York.
Emerson, J.L., D.J. Thompson, R.J. Strebing, rC. Gerbig, and V.B. Robinson. 1971.
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57 FR 24152. June 5, 1992. U.S. Emnvironmental Protection Agency (FRL-4139-7). Draft
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Equivalence of mg/kg^/day.
IARC (International Agency for Research of Cancer). 1977. IARC Monographs on the
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Khera, K.S., and W.P. McKinley. 1972s- Pre- and postnatal studies on 2,4,5-
trichlorophenooxyacetic acid, 2,4-dichlorophenoxyacetic acid and their derivatives in rats.
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August 1995
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid -11
Opresko, D. M., B. E. Sample, and G. W. Suter. 1994. Toxicological Benchmarks for
Wildlife: 1994 Revision. ES/ER/TM-86/R1.
Rehwoldt, R.E., E. Kelley, and M. Mahoney. 1977. Investigations Into the Acute Toxicity and
Some Chronic Effects of Selected Herbicides and Pesticides on Several Fresh Water Fish
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for Occupational Safety and Health, Washington, DC.
Sanderson, C.A. and L.J. Rogers. 1981. 2,4,5-Trichlorophenoxyacetic acid causes behavioral
effects in chickens at environmentally relevant doses. Science 211:593-595.
Smith, F.A., F.J. Murray, J.A. John, K.D. Nitschke, R.J. Kociba, and B.A. Schwetz. 1981.
Three-generation reproduction study of rats ingesting 2,4,5-trichlorophenoxyacetic acid in
the diet. Food and Cosmetics Toxicology 19:41-45.
Stephan, C. E, 1993. Derivation of Proposed Human Health and Wildlife Bioaccumulation
Factors for the Great Lakes Initiative. PB93-154672. Environmental Research
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Suter, G.W., M.A. Futrel, and G.A. Kerchner '992. Toxicological Benchmarks for Screening
of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak Ridge
Reservation, Oak Ridge, Tennessee. U.S. Department of Energy., Washington, D.C.
Suter n, G.W., and J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota:. 1994 Revision. ES/ER/TM-
96/R1.
Thomann, R. V. 1989. Bioaccumulation model of organic chemical distribution in aquatic
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Thomann, R. V., J. P. Connely, and T. F. Parkerton. 1992. An equilibrium model of organic
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Toxicology and Chemistry. 11:615^829:
August 1995
-------
APPENDIX B 2,4,5-Trichlorophenoxyacetic acid -12
U.S. Environmental Protection Agency. 1990. Methodology for Assessing Health Risks
Associated with Indirect Exposure to Combustor Emissions. Interim Final. Office of
Health and Environmental Assessment, Washington, D.C. January. As cited in
Pierson, T.K., A.E. Crook, S.M. Beaulieu, P.N. Graham, N.B. Jones, A.M. Reynolds, and
G.P. Vegh, 1994, Development of Human Health Based Exit Criteria for the Hazardous
Waste Identification Project, Phase IJJ Analysis.
U.S. Environmental Protection Agency. 1994. Integrated Risk Information System. July.
U.S. Environmental Protection Agency. 1993g. Wildlife Exposure Factors Handbook:
Volumes I and II. EPA/600/R-93/187a,b. Office of Science and Technology,
Washington, DC.
Yokote, M., S. Kimura, H. Kumada, and Y. Matida. 1976. Effects of some herbicides applied
in the forest to the freshwater fishes and other aquatic organisms-IV. Experiments on the
assessment of acute and.. Bull. Freshwater Fish. Res. Lab.. 26(2):85-98.
August 1995
-------
Terrestrial Toxicity - 2,4,5- .hlorophenoxyacetic acid
Cas No. 93-76-5
Chemical Nam*
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid. 2,4,5-
Trichlorophenoxy-
acetic acid. 2,4,5-
Trichlorophenoxy-
acettc acid. 2.4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5
Species
rhesus
monkey
New Zealand
white rabbits
. 1
rabbits
hamsters
hamsters
rats
Type of
Effect
tera
dvp
tera
tera
rep
tera
\
Description
NOAEL
NOAEL
NOAEL
LOAEL
LOAEL
NOAEL
Value
10
40
50
9.38
3.75
: 24
Units
mg/kg-day
mg/kg-day.
ppm
mg/ka-day
mg/kg-day
mg/kg-day
Exposure)
Route (oral.
8.C.. I.V., l.p..
Injection)
No. 5 gelatin
capsules were
administered
orally via a
stomach tube
oral
NS
oral intubation
oral intubation
gavage
Exposure
Duration/
Timing
day 22
through day
38 of
gestation
days 6- IB of
gestation
NS
days 6
through 10 of
gestation
days 6
through 10 of
gestation
days 6
through 1 5 ot
gestation
Reference
Dougherty el al.,
1976
Emerson et al.,
1971 as cited in
IAHC, 1977
Thompson et
al., 1971as cited
in Khera and
McKlnley. 1972
Collins and
Williams. 1971
Collins and
Williams, 1971
Emerson et al.,
1970 as cited in
Toxicol and
Appl Pharmac,
1970.
Comments
All ot the babies bom to date show
no evidence ot gross' leraloqenecitv
Foetal mortality and weight were
unaffected and there was no
increase in developmental variation.
Fetal mortality, the incidence of
hemorrhage in the liveborn. and the
number of malformations among the
livebom were all greatly increased.
There was an increase in the level
of embryonic mortality and the
numbers of livebom with
hemorrhages and a decrease in the
average weight per fetus. In
addition, the fetal viability per litter
was significantly decreased in a
dose-related manner.
No clinical or gross pathologic signs
were apparent in treated dams.
Liner size, number of fetal
resorptions. birth weights, and sex
ratio were unaffected.
-------
Terrestrial Toxicity - 2,4,5- Trichlorophenoxyacetic acid
Cas No. 93-76-5
Chemical Name
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acedc acid. 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic add, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acelic acid, 2,4,5-
Specles
rats
rats
rats
rats
rats
_.-
rats
mice
mice
(C57BL/6
strain)
Type of
Effect
tera
i let, dvp
V
let, dvp
rep
rep
acute
tera
tera, fet
Description
NOAEL
LOAEL
NOAEL
NOAEL
LOAEL
LD50
NOAEL
NOAEL
Value
50
11.56
5.78
3
10
300
20
46.4
Units
ppm
mg/kg-day
mg/kg-day
mo/kg-dav
mg/kg-day
ma/kg
mg/kg-diet
mg/kg-day
Exposure
Route (ore),
S.C., I.V.. l.p.,
Infection)
NS
Injection
injection
oral
oral
oral
oral
oral (stomach
intubation
suspended in a
honey sotn)
Exposure
Duration/
Timing
NS
Days 6-15.
inclusive, of
gestation
Days 6- 15,
inclusive, of
gestation
90 days for 3
generations
90 days for 3
generations
US
day 6 to 15
ot gestation
treated on •
days 6 to 14
of gestation
Reference
Thompson et
all. 1971 as cited
in Khera and
McKinley. 1972
Khera and
McKinley. 1972
Khera and
McKinley, 1972
Smith etal..
1981
Smith etal..
1981
Lehman. 1951
as cited In
Springer. 1957
Neubert and
Dillman, 1972
as cited in
Dougherty et al.,
1976
Courtney et al.,
1970
Comments
Up to a dose of 50 mg/kg-diet, no
teratogenic potential was reported in
rats.
Fetal weight, the number of dead
fetuses and the proportion ot
skeletal anomalies showed
significant differences from the
controls.
'Effects noticeable al 50mg/kg-diet
were not significant.'
No adverse effect on reproduction
was seen al this dose.
A decrease in the fertility index and
in postnatal survival were recorded
at this dosage.
A high incidence of cleft palate was
observed at doses higher than 20
ppm.
No significant increase in fetal
mortality or effect on palatal
development.
-------
Terrestrial Toxicity - 2,4,5- jhlorophenoxyacetic acid
Cas No. 93-76-5
Chemical Nam*
.
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetlc acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2.4.5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Specles
mice
(C57BU6
strain)
mice
(C57BU6
strain)
mice
(C57BL/6
strain) I
mice
(C57BL/6
strain)
mice (AKR
strain)
mice (AKR
strain)
rats
chicken
•FH" ul
Effect
(era, let
(era. let
tera. let
tera, let
tera. let
tera, let
tera
acute
Description
LOAEL
NOAEL
LOAEL
PEL
PEL
PEL
LOAEL
LD50
Value
113
21.5
113
113
113
113
4.6
53
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg
-------
Terrestrial Toxicity • 2,4,5- Trichlorophenoxyacetic acid
Cas No. 93-76-5
Chemical Nam*
' Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetlc acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acetic acid, 2,4,5-
Trichlorophenoxy-
acelic acid, 2,4,5-
Trichlorophenoxy-
acelic acid, 2,4,5-
Specle*
incubating
chicken eggs
female rats
rats
rat
mouse
dog
guinea pig
hamster
chicken
Type of
Effect
behv
behv
1
ter
acute
acute
acute
acute
acute
acute
Description
LOAEL
PEL
LOAEL
L050
L050
LD50
LD50
LD50
LD50
Value
7
6
6
300
242
100
381
425
310
Unit*
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
ma/kg
ing/kg
mfl/kg
mg^g
Exposure
Rout* (oral,
S.C., I.V., l.p.,
Inlectlon)
Injected into the
incubating eggs
•treated with
single doses'
exposure
oral
oral
oral
oral
oral
oral
Exposure
Duration/
Timing
NS
single dose
day 8 of
pregnancy
NS
NS
"NS
NS
NS
NS
Reference
. Sanderson and
Rogers, 1981 as
cited in
Crampton and
Rogers, 1982
Crampton and
Rogers, 1982
Crampton and
Rogers, 1982
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
RTECS, 1994
v Comments
'Behavioral abnormalities of
increased fear-related activity atid
learning deficits*
Behavioral abnormalities were seen
in the pups of treated mother rats.
Behavioral teratogenicity was
produced by a single exposure to
this dose.
-------
Terrestrial Toxicity - 2,4,5- jhlorophenoxyacetic acid
Cas No. 93-76-5
Chemical Name
Trichlorophenoxy-
acetic acid. 2.4.5-
Spaclaa
mammal
Type of
Eftect
acute
Description
LD50
Value
500
Units
mo/kg
Exposure .
Route (oral.
S.C., I.V., l.p.,
Inlectlon)
oral
Exposure
Duration/
Timing
NS
Reference
RTECS. 1994
Comments
-------
Freshwater Toxicity - 2,4,5-Trichlorophenoxyacetic acid
CaaNo. 93-76-5
Chemical name .
Trichlorophenoxyacetic
acid, 2.4,5-
Trichlorophenoxyacetic
acid. 2,4,5-
Trichlorophenoxyacetic
acid, 2,4.5-
Trichlorophenoxyacetic
acid. 2,4.5-
NA = Not applicable.
Species
Ceriodaphnia
dubia
Bluegill
Striped bass
Rainbow trout
,f
Type of
effect
rep
mor
mort
mor
Description
EC50
LC50
LC50
LC50
Value
17.100- 21.200
(19.055)
10.000
14.600
150-8700(1148)
Units
ug/L
ug/L
ug/L
ug/L
Test type
(static/flow
through)
NA
NA
NA
NA
Exposure
Duration/
Timing
96 hour
48 hour
96 hour
96 hour
Reference
AQUIRE. 1994
AQUIRE, 1994
AQUIRE. 1994
AQUIRE. 1994
Comments
•
-------
Freshwater Biological Uptake Meast - 2,4,5-Trichlorophenoxyacetic acid
Cas No. 93-76-5
Chemical name
Trichlorophenoxyacetic
acid. 2.4.5-
Trichlorophenoxyacettc
acid. 2,4,5-
Species
Daphnia
masne
Daphnia
magna
B-factor
(BCF, BAF,
BMP)
BCF
BCF
Valu«
70-1CIO
(84.14)
10-16
(13.39)
Measured or
Predicted
(m.P)
m
m
Units
NS
NS
Reference
Isensee. 1976 as
cited in AQUIRE,
1994
Yockimetal.,
1978 as cited in
AQUIRE. 1994
Comments
Life stage not
reported.
Life stage not
reported.
NS - Not specified
-------
Freshwater Biological Uptake Measures - 2,4,5-Trichlorophenoxyacetic acid
Cas No. 93-76-5
Chemical name
Trichlorophenoxyacetic
acid, 2.4,5-
Spflclaa
plants
B-factor
(BCF, BAF,
BMP)
BCF
Value
1
Measured or
Predicted
(m.p)
-
Units
NS
Reference
U.S. EPA. 1990e
Comments
NS = Not specified.
-------
APPENDIX B i Vanadium - 1
lexicological Profile for Selected Ecological Receptors
Vanadium
Cas No.: 7440-62-2
Summary: This profile on vanadium summarizes the lexicological benchmarks and
biological uptake measures (i.e., bioconcentration, bioaccumulation, and biomagnification
factors) for birds, mammals and fish representing the generic freshwater and terrestrial
ecosystems. Toxicological benchmarks were derived for developmental, reproductive or other
effects reasonably assumed to impair population growth and survival. Bioconcentration
factors (BCFs), bioaccumulation factors (BAFs) and, if available, biomagnification factors
(BMFs) are also summarized for the ecological receptors, although BAFs for the freshwater
ecosystem were calculated for organic constituents with log Kow between 5 and 6.5. For the
terrestrial ecosystem, these biological uptake measures also include terrestrial invertebrates
(i.e. insects and earthworms). In addition, the entire lexicological data base compiled during
this effort is presented at the end of this profile and includes additional studies and existing
regulatory benchmarks (eig., National Ambient Water Qualily Criteria or NAWQC). This
profile represents the most current information and may differ from the data presented in the
technical support document for the Hazardous Waste Identification Rule (HWIR): Risk
Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presenls ihe rationale behind lexicological benchmarks used to derive protective
media concentrations (Cpro) for the generic freshwater ecosystem. Table 1 contains
benchmarks for mammals and birds associated with the freshwater ecosystem and Table 2
contains benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including
aquatic plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: A number of toxicily sludies were identified that focused on the effecis of
vanadium on laboratory rats and mice. Paternain et al. (1990) administered vanadyl sulfale
peniahydrale at doses of 37.5, 75, and 150 mg/kg-day by gavage to Swiss mice on days 6-15
of pregnancy. Maternal toxicity, embryoioxiciiy, feioioxicity and teraiogenicily were
observed al dosages as low as 37.5 mg/kg-day. Anoiher chronic siudy was identified in
which female Sprague-Dawley albino rats were fed sodium metavanadate at doses of 5, 10, or
20 mg/kg-day, 14 days prior to mating with males who had been previously treated for 60
days, during gestation, as well as, 21 days following delivery of the pups (Domingo et al.,
1986). At the 5 mg/kg-day dosage, the body weight and length of ihe ral pups nursed by
vanadium-treaied mothers was significantly lower than the controls. Bosque et al. (1993)
reported a significant decrease in the felal body weight per litier of albino Swiss mice wilh a
single intra-peritoneal injection of 25 mg sodium metavanadate/kg on gestation day 12.
August 1995
-------
APPENDIX B Vanadium - 2
Injections were administered on one of days 9-12 to determine whether embryotoxicity and
fetotoxicity varied with the day of exposure.
The Bosque et al. (1993) study was not considered suitable for the derivation of a mammalian
benchmark because the dose was administered intra-peritoneally and extrapolation. from the
injection route of exposure to typical wildlife exposure routes (i.e. oral) would increase the
uncertainty associated with the resulting benchmark. Although both the Patemain et al.
(1990) study and the Domingo et al. (1986) study investigated effects associated with orally
administered doses of vanadium, the Domingo et al. (1986) study was preferred since the test
species employed, the rat, has been identified as being more susceptible to vanadium than the
mouse, the test species used in the Paternain et al. (1990) study (Pham-Huu-chanh, 1965).
The Domingo et al, (1986) study was selected for the derivation of lexicological benchmarks
because it 1) illustrated clear dose-response data, 2) studied reproductive endpoints, 3)
demonstrated an orally administered dosage of vanadium and 4) investigated the toxicity
effects of vanadium on a particularly sensitive test species. Domingo et al. (1986) reported a
LOAEL of 5 mg/kg-day at which significant decreases were observed in the development of
pups in vanadium-treated groups. The selected study LOAEL was divided by 10 to provide a
LOAEL-NOAEL safety factor. The LOAEL/10 from Domingo et al. (1986) was then scaled
for species that were representative of a freshwater ecosystem using a cross-species scaling
algorithm adapted from Opresko et al. (1994):
Benchmark = NOAEL. x
where NOAELj is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight
of the wildlife species, and BW, is the body weight of the test species. This is the same
default methodology EPA provided for carcinogenicity assessments and reportable quantity
documents for adjusting animal data to an equivalent human dose (57 FR 24152). Since the
Domingo et al. (1986) study documented reproductive effects from exposure to female rats,
the female body weights of the representative species were used in the scaling algorithm to
obtain lexicological benchmarks. Based on the data set for vanadium and as the Domingo et
al. (1986) study provided a LOAEL rather than a NOAEL, the benchmarks developed from
this study were categorized as provisional.
Birds: There was only one study identified which investigated the effects of vanadium
toxicity in avian species. Romoser et al. (1961) fed 7-day chicks a diet containing vanadium
as a calcium salt from days 7 through 28. A depression in the rate of weight gain was
observed above 20 ppm. These results suggest a NOAEL of 20 ppm. No information on
daily food consumption rates were provided therefore, the use of an allpmetric equation was
required to convert doses from dietary ppm to mg/kg-day:
Food consumption = 0.075( W0'8449 ) where W is body weight in kg (Nagy, 1987 ).
August 1995
-------
APPENDIX B Vanadium . 3
The geomean of the body weight of 1 week and 4 week old Vantress x Arbor Acre male
chicks was determined to be 0.487 kg (Parkhurst, 1995). The food consumption rate which
was estimated as being 0.041 kg/day was multiplied by the dietary ppm value and divided by
the body weight. In this way, the daily dose was determined to be 1.68 mg/kg-day. The
value was then scaled for species representative of a freshwater ecosystem using the cross-
species scaling algorithm adapted from Opresko et al. (1994). Since the Romoser et al.
(1961) study documented effects of vanadium exposure to male chicks, mean male body
weights of the representative species were used in the scaling algorithm to obtain the
toxicological benchmarks. Based on the data set for vanadium and since the Romoser et al.
(1961) study provided a NOAEL, the benchmarks developed from the study were categorized
as adequate.
Fish and aquatic invertebrates: No AWQC or Final Chronic Value (FCV) was available for
vanadium. Therefore, a Secondary Chronic Value (SCV) of 1.9 E-02 mg/1 (Suter and
Mabrey, 1994) was utilized. Because, an SCV was utilized, the benchmark was categorized
as interim.
Aquatic Plants: The benchmarks for aquatic plants were either: (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEC) for vascular aquatic
plants (e.g., duckweed) or 2) an effective concentration (ECXX) for a species of freshwater
algae, frequently a species of green algae (e.g., Selenastrum capricornutum). No CV was
reported for vanadium and, therefore, no benchmark was developed. As described in Section
4.3.6, all benchmarks for aquatic plants were designated as interim.
Benthic community: The vanadium benchmark protective of benthic organisms is pending a
U.S. EPA review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Vanadium - 4
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
FUpfBWaUtfvw
Spaefw
mink
river otter
bald eagle
osprey
great blue heron
mallard
tester scaup
spotted sandpiper
herring gut
kingfisher
BwchOWdC
Value* «8*t"
*
aquatic •
organisms
-
-
Original
VaUi*
mgtl
1.9E-02
-
Description
scv
-
-
OrJgiMl Souro*
Suter & Mabrey,
1994
•
-
'Benchmark Category, a • adequate, p « provisional, i »interim; ID * insufficient data; a (') indicates (hat the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Vanadium - 5
II. Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains
benchmarks for mammals, birds, plants and soil invertebrates representing the generic
terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: As mentioned previously in the freshwater ecosystem discussion, several toxicity
studies were identified that focused on the effects of vanadium on laboratory mammals.
Since no additional studies for terrestrial mammals were found, the same surrogate study
(Domingo et al., 1986) was used to calculate benchmark values for mammalian species
representing the jgeneral terrestrial ecosystem. The LOAEL from the Domingo et al. (1986)
study was scaled for species in the terrestrial ecosystem using the cross-species scaling
algorithm adapted from Opresko et al. (1994). Since the Domingo et al. (1986) study
documented reproductive effects from vanadium exposure to female rats, female body weights
for each representative species were used in the scaling algorithm to obtain lexicological
benchmarks. Because the benchmarks developed from the Domingo et al. (1986) study
required the use of a safety factor to extrapolate from the NOAEL to LOAEL, they were
categorized as provisional, as in the aquatic ecosystem.
Birds: No additional avian toxicity studies were identified for species representing the
terrestrial ecosystem. Thus, for avian species in ihe terrestrial ecosystem, the NOAEL of 1.68
mg/kg-day from the Romoser et al. (1961) study was used as the benchmark value. This
value was then scaled for terrestrial species using the cross-species scaling algorithm adapted
from Opresko. et al. (1994). Based on the avian data set for vanadium, the benchmarks
developed from the Romoser et al. (1961) study were categorized as adequate.
Plants: Adverse effects, levels for terrestrial plants were identified for endpoints ranging from
percent yield to root length. As presented in Will and Suter (1994), phytotoxicity
benchmarks, were selected by rank ordering the Lowest Observable Effects Concentration
(LOEC) values and then approximating the 10th percentile. If there were 10 values, the 10th
percentile LOEC was used. Such LO' Cs applied to reductions in plant growth, yield
reductions, or other effects reasonably "assumed to impair ihe ability of a plant population to
sustain itself, such as a reduction in seed elongation. The benchmark for terrestrial plants was
2 mg/kg based on the lowest LOEC presented by Will and Suier (1994). As less than 10
studies were presented by Will and Suter (1994), the phytoioxicity benchmark of 2 mg/kg was
categorized as interim.
Soil community: Adequate data with which to derive a benchmark protective of the soil
community were not identified.
August 1995
-------
APPENDIX B
Vanadium - 6
Table 3. lexicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
ffcprwerrfwJYB
Sped**
daw mouse
short-tailed
shrew
meadow vote
Eastern
cottontail
red fox
raccoon
white-tailed
-------
APPENDIX B
Vanadium • 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive
protective surface water and soil concentrations for constituents considered to bioconcentrate
and/or bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake
values and sources are presented in Table 4 for ecological receptor categories: fish in the
limnetic or littoral ecosystem, aquatic invertebrates, earthworms, other soil invertebrates,
terrestrial vertebrates, and plants. For metals, BCFs are whole-body bioconcentration factors
and refer to total surface water concentrations (versus freely dissolved concentrations).
Consequently, all calculations of acceptable tissue concentrations (TC) represent whole-body
concentrations. The following discussion describes the rationale for selecting the biological
uptake factors and provides the context for interpreting the biological uptake values.
Insufficient data were identified to determine the whole-body BCF for silver in fish, aquatic
invertebrates, terrestrial vertebrates and earthworms. A whole plant BCF value of 5.5 E-03
was derived from U.S. EPA. (1992e). For metals, empirical data were used to derive the
BCF for aboveground forage grasses and leafy vegetables. In particular, the uptake response
slope for forage grasses was used as the BCF for plants in the terrestrial ecosystem since
most of the representative plant-eating species feed on wild grasses.
Table 4. Biological Uptake Properties
•OOfogiOftl
: r»c«ptar
fish
littoral
trophic level 2
invertebrate*
terrestrial
vertebrates
terrestrial
invertebrates
earthworms
plants
BCF,BAF,or
BSAF
-
•
-
BCF
ifpid.i>«a«i ot
whohhbody
•
•
•
-
whole-plant
vafu*
10
ID
10
' ID
ID
5.5E-03
•cure*
<
-
•
U.S.EPA, 1992e
d s refers to dissolved surface water concentration
t * refers to total surface water concentration
ID = refers to insufficient data
August 1995
-------
APPENDIX B Vanadium-8
References
AQUIRE (AOUztic Toxicity_/nformation /?Etrieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Bosque, M.A, J.L. Domingo, J.M. Llobet and J. Corbella. 1993. Variability in the '
embryotoxicity and fetotoxicity of vanadate with the day of exposure. Vet Hum.
Toxicol. 35.
Carlton, B.D., M.B. Beneke and G.L Fisher. Assessment of the teratogenicity of ammonium
vanadate using Syrian Golden hamsters. Env. Research 29: 256-262.
Clement Associates, Inc. 1990. Draft: Toxicological Profile for Vanadium and Compounds.
Prepared for Agency for Toxic Substances and Disease Registry (ATSDR), U.S.
Public Health Service.
i
Domingo, J.L, J.M. Llobet and J.M. Tomas. 1985. Short-term toxicity studies of vanadium in
rats. J. Appl. Toxicol., V. 5, No. 6.
Domingo, J.L, J.L Paterriain, J.M Llobet and J. Corbella. 1986. Effects of vanadium on
reproduction, gestation, parturition, and lactation in rats upon oral administration. Life
Sciences, 39: 819-824.
Domingo, J.L., J.M. Llobet, J.M Tomas and J. Corbella. 1986a. Influence of chelating
agents on the toxicity distribution and excretion of vanadium in mice. /. Appl. Tox.
V6 (5) 337-341.
Domingo, J.L. 1994. Metal-induced developmental toxicity in mammals: a review, /. Toxicol.
and Env. Health., 42:123-141.
57 FR 24152. June 5, 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg3/4/day.
Hamilton, S.J. and K.J. Buhl. 1990. Safety assessment of selected inorganic elements to fry of
chinook salmon (Oncorhynchus tshawytschd). Ecotox. and Env. Safety 20: 307-324.
Kowalska, M. 1988. The effect of vanadium on lung collagen content and composition in
two successive generations of rats. Toxicol Lett 41:203-208. As cited in Clement
Associates, Inc. 1990. Draft: Toxicological Profile for Vanadium and Compounds.
Prepared for Agency for Toxic Substances and Disease Registry (ATSDR), U.S. Public
Health Service.
August 1995
-------
APPENDIX B Vanadium - 9
Luckey, T.D. and B. Venugopal. Metal toxicity in mammals (1): Physiologic and chemical
basis for metal toxicity. Plenum Press, N.Y.
Marmo, E., M.G. Matera, R. Acamora, C. Vacca, D. Desantis, S. Maione, V. Susanna, S.
Chieppa, V. Guarino, R. Servodio, B. Cuparencu, and F. Rossi. 1987. Prenatal and
postnatal metal exposure: effect on vasomotor reactivity development of pups. Current
therapeutic research 42 (4).
Nagy, K.A. 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol. Mono. 57:111-128.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. ToxicologicalBenchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S. Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
Parkhurst, C.R. 1995. Personal communication. Department of Poultry Science, North
Carolina State University, Professor.
Paternain, J.L., J.L. Domingo, M. Gomez, A. Ortega, and J. Corbella. 1990. Developmental
toxicity of vanadium in mice after oral administration. /. Appl. Tox.,V. 10(3): 181-186.
Pham-Huu-chanh. 1965. The comparitive toxicity of sodium chromate molybdate, tungstate,
and metavanadate, Arch. Int. Pharmacodyn. 154:243.
Ridgway, L.P and D.A. Kamofsky. 1952. The effects of metals on the chick embryo: Toxicity
and production of abnormalities in development. Ann. N.Y. Acad. Sci. 55:203.
Romoser, G.L, W.A. Dudley, L.J. Machlin and L. Loveless. 1961. Toxicity of vanadium and
chromium for the growing chick. Poultry Sci. V.40: 1171-1173.
Schroeder, H.A., J.J. Balassa, I.H. Tipton. 1963. Abnormal trace metals in man- vanadium.
Journal of Chronic Disease. 16:1047-1071. As cited in Clement Associates, Inc. 1990.
Draft: Toxicological Profile for Vanadium and Compounds. Prepared for Agency for
Toxic Substances and.Disease Registry (ATSDR), U.S. Public Health Service.
Schroeder, H.A., M. Mitchener and A.P. Nason. 1970. Zirconium, niobium, antimony,
vanadium and lead in rats: Life term studies. J. Nutrition, 100:59-68.
Schroeder, H.A., M. Mitchener. 1975. Life-time effects of mercury, methyl mercury, and nine
other trace metals on mice. /. Nutr. 105:245-252. As cited in Clement Associates, Inc.
1990. Draft: Toxicological Profile for Vanadium and Compounds. Prepared for Agency
for Toxic Substances and Disease Registry (ATSDR), U.S. Public Health Service.
August 1995
-------
Terrestrial Toxicity - Vanadium
Cas No. 7440-62-2
Chemical
Name
vanadium
sodium
metavanadate
vanadyl sulfate
jentahydrate
sodium
metavanadate
vanadium
vanadium
vanadium
vanadium
vanadium
vanadium
Species
rat
rat
mice
mice
chick
rat
rat
mouse
mouse
rat
Type of '
Effect
rep
dev
dev
dev
growth
dev
path
path
path .
chronic
Description
NOAEL
LOAEL
LOAEL
PEL
NOAEL
LOAEL
NOAEL
NOAEL
NOAEL
NOEL
Value
8.4
5
37.5
25
1.68 :
2.8
0.7
4.1
6.54
0.9
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg
mg/kg-day
mg/kg-day
mg/kg;day
mg/kg-day
mg/kg-day
mg/kg-day
Exposure
Route (oral,
s.c., i.v., l.p.,
injection)
gavage
gavage
gavage .
'P
oral
oral
oral
oral
oral
oral
Exposure Duration
/Timing
60 days
60 days
9 days
1 dose
3 Weeks
2 generations
°
2.5 years
2 years
2.5 years
Weaning until natural
death
Reference
Domingo et at.. 1986
as cited in ATSDR, •
1992
Domingo et al., 1986
Paternain et al., 1990
Bosqueetal, 1993
Romoser et al, 1961
Kowalska et al.. 1988
as cited in ATSDR,
1992
Schroeder et al., '
1970 as cited in
ATSDR
Schroeder and
Balassa, 1967 as
cited jn ATSDR. 1992
Schroeder and
Mitchener, 1 975 as
cited in ATSDR, 1992
Schroeder et al.,
1970
Comments
Reduced pup weight and
length.
Maternal toxicity,
embryotoxicity. (eloloxicity
and teratogenicity.
Decreased fetal body
weight and length
Depression in rate of weight
gain
Altered lung collagen in
pups of adults exposed to
vanadium over a lifetime.
Dose converted from ppm
by calculating dose from the
food and the dose from the
drinking water and adding .
to get a single dose level in
mg/kg-day.
-------
APPENDIX B Vanadium - 10
Suter D, G.W., M.A. Futrell, and G.A. Kerchner.. 1992. Toxicological Benchmarks for
Screening of Potential Contaminants of Concern for Effects on Aquatic Biota on the Oak
Ridge Reservation, Oak Ridge, Tennessee. DE93-000719. Office of Environmental
Restoration and Waste Management, U.S. Department of Energy, Washington, DC.
Suter n, G.W., J.B. Mabrey. 1994. Toxicological Benchmarks for Screening Potential
Contaminants of Concern for Effects on Aquatic Biota: 1994 Revision. ES/ER/TM-96/R1.
U.S. Department of Energy, Oak Ridge National Laboratory, Oak Ridge, TN
U.S. EPA (Environmental Protection Agency). 1980. A Screening procedure for the
Impacts of Air Pollution Sources on Plants, Soils, and Animals. EPA 450/2-81-078.
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1988. Recommendations for and
Documentation of Biological Values for Use in Risk Assessment. P338-179874.
Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1992e. Technical Support Document for Land
Application of Sewage Sludge, Volume I and 11. EPA 822/R^93-001a. Office of Water,
Washington, DC.
Venugopal, B. and T.D. Luckey. Metal toxicity in mammals (2); Chemical toxicity of metals
and metalloids. Plenum Press, N.Y., 1978.
Wide, M. 1982. Effect of short-term exposure to five industrial metals on the embryonic and
fetal development of the mouse. Environmental research 33, 47-53.
Will, M.E and G.W. Suter II. 1994. Toxicological Benchmarks for Screening of Potential
Contaminants of Concern for Effects on Terrestrial Plants: 1994 Revision. DE-AC05-
84OR21400. Office of Environmental Restoration and Waste Management, U.S.
Department of Energy, Washington, DC.
August 1995
-------
Freshwater Toxicity - Vanadium
Cas No. 7440-62-2
Chemical
Name
Vanadium
Vanadium
Vanadium
Vanadium
Vanadium .
Species
aquatic
organisms
fish
daphnid
(ish
daphnid
Type of
Effect
chronic
chronic
chronic
chronic
chronic
Description
NAWQC
CV
CV
EC20
EC20
Value
3
80
>940
41
430
Units
ug/L
ug/L
ug/L
ug/L
ug/L
Test Type
(Static/Row
Through)
NS
MS
NS
NS
NS
Exposure
Duration
/Timing
NS
NS
NS
NS
NS
Reference
Suteretal.. 1992
Suteretal., 1992
Suteretal.. 1992
Suteretal., 1992
Suteretal.. 1992
Comments
-------
Terrestrial To~ /-Vanadium
Cas No. 7440-62-2
Chemical
Name
vanadium
Species
mouse
Type of
Effect
let
Description
AEL
Value
1
Units
mM
Exposure
Route (oral,
s.c., i.v., i.p..
injection)
i.v.
Exposure Duration
/Timing
1
Reference
Wide, 1984
Comments
Decreases in fetuses with
mature skeletons.
i
-------
Terrestrial Biological Uptake Measures - Vanadium
Cos No. 7440-62-2
Chemical
Name
vanadium
Species
plant
B-factor
(BCF, BAF.
BMP)
BCF
Value
0.0055
Measured
or
Predicted
. (m-P)
P
units
(ug/g DW plant)/(ug/g
soil)
Reference
U.S. EPA, 1990e
Comments-
-------
Freshwater Biological Up j Measures - Vanadium
Cas No. 7440-62-2
Chemical
Name
Species
B-factor
(BCF, BAF,
BMP)
Value
Measured
or
Predicted
(m,p)
Units
Reference
.Comments
-------
APPENDIX B Zinc - 1
lexicological Profile for Selected Ecological Receptors
Zinc
CasNo.: 7440-66-6
Summary: This profile on zinc summarizes the lexicological benchmarks and biological uptake
measures (i.e., bioconcentration, bioaccumulation, and biomagnification factors) for birds,
mammals, daphnids and fish, aquatic plants and benthic organisms representing the generic
freshwater ecosystem and birds, mammals, plants, and soil invertebrates in the generic terrestrial
ecosystem. Toxicological benchmarks for birds and mammals were derived for developmental,
reproductive or other effects reasonably assumed to impact population sustainability. Benchmarks
for daphnids, benthic organisms, and fish were generally adopted from existing regulatory
benchmarks (i.e., Ambient Water Quality Criteria). Bioconcentration factors (BCFs),
bioaccumulation factors (BAFs) and, if available, biomagnification factors (BMFs) are also
summarized for the ecological receptors, although some BAFs for the freshwater ecosystem were
calculated for organic constituents with log Kow between 4 and 6.5. For the terrestrial ecosystem,
these biological uptake measures also include terrestrial vertebrates and invertebrates (e.g.,
earthworms). The entire lexicological data base compiled during this effort is presented at the
end of this profile. This profile represents the most current information and may differ from the
data presented in the technical support document for the Hazardous Waste Identification Rule
(HWIR): Risk Assessment for Human and Ecological Receptors.
I. Toxicological Benchmarks for Representative Species in the Generic Freshwater
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (€_) for the generic freshwater ecosystem. Table 1 contains benchmarks
for mammals, and birds associated with the freshwater ecosystem and Table 2 contains
benchmarks for aquatic organisms in the limnetic and littoral ecosystems, including aquatic
plants, fish, invertebrates and benthic organisms.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Three studies were identified which investigated the effects of zinc exposure on
mammals. Samanta and Pal (1985) studied the effects of 4000 ppm of zinc fed to male rats.
As the authors did not provide daily food consumption rates, an allometric equation was utilized:
Food consumption - 0.056(W°-6611) where W is body weight in kg ( Nagy, 1987 ).
Using the reported body weight of 0.162 kg, 0.017 mg/kg-day was estimated as the daily dietary
intake of zinc. After 32 days of exposure at this dose level, male rats exhibited decreased sperm
motility and reduced fertilizing capacity. In another study, Bleavins et al. (1983) found that an
oral dose of 21 mg Zn/kg-day fed to female mink for 25 weeks had no effect on the length of
the gestation .period or litter size. A NOAEL of 21 mg/kg-day was inferred from these results
August 1995
-------
APPENDIX B Zinc - 2
for reproductive effects. Schlicker and Cox (1968) observed an increased percentage of fetal
resorptions in female rats fed a 0.4% zinc oxide-amended diet. The rats were fed a 0%, 0.2%
or 0.4% zinc diet for 21 days prior to mating and up until a fetal age of 15 days. As the
quantity of food consumed was not included in the study, the allometric equation presented above
(Nagy, 1987) was utilized to estimate the daily dose of dietary zinc. The geomean (0.174 kg) of
the reported body weight of the test species, the food consumption rate of 0.0176 kg/day and the
percentage of zinc oxide in the diet, were used to derive a NOAEL of 202.4 mg/kg-day.
Although the Samanta & Pal (1985) study measures reproductive endpoints that could impair a
wildlife population's sustainability, the short duration of the study and its failure to establish a
dose-response relationship made it unsuitable for the calculation of a benchmark value. The
Bleavins et al. (1983) study focused on the effects of dietary zinc at a single dose and therefore,
an adequate dose-response relationship could not be established. The Schlicker and Cox (1968)
study (1) focused on a reproductive endpoint, and (2) established an adequate dose-response
relationship and for these reasons the NOAEL of 202 mg/kg-day was chosen for derivation of
a benchmark value. This value was scaled for species representative of a "fresh water ecosystem
using a cross-species scaling algorithm adapted from Opresko et al. (1994):
Benchmark^ = NOAEL, x
where NOAELj is the NOAEL (or LOAEL/10) for the test species, BWW is the body weight of
the wildlife species, and BW, is the body weight of the test species. This is the same default
methodology EPA provided for carcinogenicity assessments and reportable quantity documents
for adjusting animal data to an equivalent human dose (57 FR 24152). Since the Schlicker and
Cox (1968) study documented reproductive effects on female rats, female body weights for each
representative species were used in the scaling algorithm to obtain the lexicological benchmarks.
Based on the data set for zinc, the benchmarks developed from the Schlicker and Cox (1968)
study were categorized as adequate with an "*" to denote that adverse effects in mammals may
occur at the benchmark level.
Birds: No suitable studies were identified which investigated reproductive or developmental
toxicity of zinc in avian species.
Fish and aquatic invertebrates: The Final Chronic Value (FCV) for zinc of 1.1 E-Olmg/1 was
selected as the benchmark protective of fish and aquatic invertebrates (Suter and Mabrey, 1994).
The FCV for zinc is a function of water hardness and is calculated using the equation
e(0.8473[ino.76i4) ^ ^ water hardness variable normalized to 100 mg/1. Since the
FCV was derived from the AWQC document, the benchmark was categorized as adequate, with
a "*" to indicate that adverse effects to aquatic organisms may occur at the benchmark level.
Aquatic Plants: The benchmarks for aquatic plants were either (1) a no observed effects
concentration (NOEC) or a lowest observed effects concentration (LOEQ for vascular aquatic
plants (e.g., duckweed) or (2) an effective concentration (ECXX) for a species of freshwater algae,
August 1995
-------
APPENDIX B Zinc - 3
frequently a species of green algae (e.g., Selenastrum capricornunuri). The aquatic plant
benchmark for zinc is 0.030 mg/1 based on the incipient inhibition of growth in Selenastrwn
capricornutum. As described in Section 4.3.6, all benchmarks for aquatic plants were designated
as interim.
Benthic community: The zinc benchmark protective of benthic organisms is pending a U.S. EPA
review of the acid volatile sulfide (AVS) methodology proposed for metals.
August 1995
-------
APPENDIX B
Zinc - 4
Table 1. lexicological Benchmarks for Representative Mammals and Birds
Associated with Freshwater Ecosystem
R*pr*Mai*tfy»
•QptAtftst
mink
river oUBf
bald eagle
osprey
.great blue heron
mallard
lesser scaup
spotted sandpiper
herring gut
kingfisher
8*nchiR*rfc
Van***?*?.
day
142.B (a*)
79.53 (a')
ID
ID
ID
ID
ID
ID
ID
ID
Study
ttp+clo
rat
rat
-
-
-
-
-
-
8
Ortgfoal SOUK*
Schfcker and Cox,
1968
SchScker and Cox,
1968
-
:
• •
•
•
•
•Benchmark Category, a « adequate, p « provisional, i * interim, ID = insufficient data; a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
Table 2. Toxicological Benchmarks for Representative Fish
Associated with Freshwater Ecosystem
ttepr»**ftt»tfve
fish and aquatic
invertebrates
aquatic plants
benlhic community
fienchmaric
V$fe**
mgfL
1.1 E-01 (a')
0.030
under review
Study
$pe$!iBe
aquatic
organisms
Sftanastrvm
capricomutum
Original
V*fc»
«0rl
1.1 E-01
O.OX
Oe«od|>Bo»
FCV
-
'
OHfi^tal^ourae
AWQCTabte
Suter & Mafarey.
1994
•
•Benchmark Category, a = adequate, p = provisional, i = interim, ID = insufficient data; a (*) indicates that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Zinc-5
IL Toxicological Benchmarks for Representative Species in the Generic Terrestrial
Ecosystem
This section presents the rationale behind lexicological benchmarks used to derive protective
media concentrations (C ) for the generic terrestrial ecosystem. Table 3 contains benchmarks
for mammals, birds, plants and soil invertebrates representing the generic terrestrial ecosystem.
Study Selection and Calculation of Toxicological Benchmarks
Mammals: Because no additional mammalian toxicity data were identified, the Schlicker and
Cox (1968) study used to calculate a freshwater mammalian benchmark was also used for the
terrestrial ecosystem. The NOAEL from Schlicker and Cox (1968) was scaled for species in the
terrestrial ecosystem using the cross-species scaling algorithm adapted from Opresko et al.,
(1994), presented above. Since the Schlicker and Cox (1968) study documented reproductive
effects from exposure to zinc in female rats, female body weights for each representative species
were used in the scaling algorithm to obtain the lexicological benchmarks. Based on the data
set for zinc, the benchmarks developed from the Schlicker and Cox (1968) study were
categorized as adequate with a "*" to indicate thai adverse effects may occur at the benchmark
level.
Birds: As mentioned in ihe freshwater ecosystem discussion, adequate data with which to derive
a benchmark protective of the avian community were not identified.
Plants: Adverse effects levels for terrestrial plants were identified for endpoinls ranging from
perceni yield to root length. As presented in Will and Suter (1994), phytotoxicity benchmarks,
were selected by rank ordering the Lowest Observable Effects Concentration (LOEC) value:; and
then approximating the 10th percentile. If there were 10 values, the 10th percentile LOEC was
used. Such LQECs applied to reductions in plant growth, yield reductions, or other effects
reasonably assumed to impair the ability of a plant population to sustain itself, such as a
reduction in seed elongation. The benchmark for terrestrial plants was 50 mg/kg as this is the
10th percentile LOEC presented by Will and Suter (1994). The phytotoxicity benchmark was
categorized as provisional as there were more than 10 values presented by Will and Suter (1994).
Soil Community: For the soil community, the lexicological benchmarks were established based
on methods developed by the Dutch National Institute of Public Health and Environmental
Protection (RIVM). In brief, the RFVM approach estimates a concentration at which the no
observed effect concentration (NOEC) for 95% of the species within the community is; not
exceeded. A minimum data set was established in which key structural and functional
components of the soil community (e.g. decomposer guilds, grazing guilds) encompassing
different sizes of organisms (e.g., microfauna, mesofauna, and macrofauna) were represented.
Measurement endpoints included reproductive effecls as well as measures of mortality, growih,
and survival. The derived zinc benchmark deemed protective of ihe soil community is 3.6 E-02
mg/kg. Since ihe zinc daia sel coniains NOECs and/or LOECs for al least four of ihe
represeniative species oudined in ihe minimum soil daia set, the soil community benchmark is
categorized as interim.
August 1995
-------
.APPENDIX B
Zinc • 6
Table 3. Toxicological Benchmarks for Representative Mammals and Birds
Associated with Terrestrial Ecosystem
WeflWieefWIRflfle'
flpeotee
doer mouse
short-tailed
shrew
meadow vole
Eastern
cottontail
red fox
raccoon
white-tailed deer
red- tailed hawk
American kestrel
Northern
bobwhite
American robin
American
woodcock
plants
soil community
8*tM9tMMtlC ;
: V«faNft \
• «w/fcHnr !
352.30 (a*)
362.23 (a*)
2.94.33 (a*)
124.35 (a*)
92.29 (a*)
88.81 (a*)
44.30 (a*)
ID
ID
ID
ID
ID
50 mg/Vg (p)
3.6 E-02 (i)
mg/Vg
3t»Kfy
ffperje*
rat
rat
rat
rat
rat
rat
rat
-
-
-
-
terrestrial
plants
soil
1 invertebrates
>
CHect
rep
rep
rep
rep
rep
rep
rep
-
-
-
-
growth/
yield
chronic
m*fy
V«fa»
mgfo&f
2.02 E+02
2.02 E+02
2.02 E+02
2.02 E+02
2.02 Ei-02
2.02 E+02
2.02 E+02
•
-
•
'•
50mg/kg
3.6 E-02 (i)
mg/kg
> '
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
NOAEL
-
-
-
- ' •
10* percentile
LOEC
NOEC
8F
-
-
-
•
•
•
-
-
-
-
-
-
-
0^,^
Schlicker & Cox,
1968
Schlicker & Cox,
1968
Schlicker & Cox,
1968
Schlicker & Cox,
1968
Schlicker & Cox,
1968.
Schlicker & Cox,
1968
Schlicker & Cox,
1968
•
•
-
- .
Will and Suter,
1994
Aldanberg and
Slob, 1993
•Benchmark Category, a - adequate, p = provisional, i = interim, ID = insufficient data; a (*) wKfcales that the benchmark
value was an order of magnitude or more above the NEL or LEL for other adverse effects.
August 1995
-------
APPENDIX B Zinc - 7
in. Biological Uptake Measures
This section presents biological uptake measures (e.g., BCFs, and BAFs) used to derive protective
surface water and soil concentrations for constituents considered to bioconcentrate and/or
bioaccumulate in the generic aquatic and terrestrial ecosystems. Biological uptake values and
sources are presented in Table 4 for ecological receptor categories: fish in the limnetic or littoral
ecosystem, aquatic invertebrates, earthworms, other soil invertebrates, terrestrial vertebrates, and
plants. For metals, BCFs are whole-body bioconcentration factors and refer to total surface water
concentrations (versus freely dissolved concentrations). Consequently, all calculations of
acceptable tissue concentrations (TC) represent whole-body concentrations. The following
discussion describes the rationale for selecting the biological uptake factors and provides the
context for interpreting the biological uptake values.
The whole-body fish BCF value for zinc was a geometric mean of 2 values presented by
Deutch et al., 1980. A value of 130 was measured in 3-spine stickleback and a value of 200 was
derived for 15-spine stickleback. The geometric mean of 161 was therefore utilized as the BCF
value for fish. BCF values for muscle were not included because ecological receptors are likely
to eat the whole fish or, in the least, will not necessarily distinguish between the fillet and other
parts of the fish. Insufficient data were identified to determine the BCF value in aquatic
invertebrates, terrestrial vertebrates and terrestrial invertebrates. The bioconcentration factor for
earthworms was derived from the geomean of two BCFs for earthworms. Davies (1983)
measured a BCF range in earthworms of 0.68-5.4. Helmke (1979) presented a BCF range of
2-3 for earthworms. A BCF of 2.2 was therefore presented for worms. A whole plant BCF value
of 9.6 E-02 was derived from U.S. EPA (1992e). For metals, empirical data were used to derive
the BCF for aboveground forage grasses and leafy vegetables. In particular, the uptake response
slope for forage grasses was used as the BCF for plants in the terrestrial ecosystem since most
of the representative plant-eating species feed on wild grasses.
August 1995
-------
APPENDIX B
Zinc • 8
Table 4. Biological Uptake Properties
•oofegkai
raoeptor
fish
littoral
trophic tovd 2
invertebrates
terrestrial
vertebrates
terrestrial
invertebrates
'earthworm*
plant*
BCF, BAF,<*
83 AF
BCF
•
BCF
BCF
tiptdbMiedor
whole-body
•
•
whole-body
whole-plant
v*u*
161
ID
ID
ID
2.2E+00
9.6 E-02
MHtfO*
Deutch et al.. 1980 as cited in
AQUIRE
-
. • -.
-
Davies. 1983; Helmke, 1979
U.S. EPA, 1992e
d » reters to dissolved surface water concentration
t - reters to total surface water concentration
ID - reters to insufficient data
August 1995
-------
APPENDIX B Zinc-9
References
AQUIRE (AQUatic Toxicity Information REtrieval Database). 1995. Environmental Research
Laboratory, Office of Research and Development, U.S. Environmental Protection Agency,
Duluth, MN.
Aldenberg, T. and W. Slob. 1993. Confidence limits for hazardous concentrations based on
logistically distributed NOEC toxicity data. Ecotoxicology and Environmental Safety.
25:48-63.
Ansari, M. S., W. J. Miller, M. W. Neathery, J. W. Lassiter, R. P. Gentry, and R. L. Kinciid.
1976. Zinc metabolism and homeostasis in rats fed a wide range of high dietary zinc
levels. Proc. Soc. Exp. Biol. Med. 152:192-194.
Aulerich, R. J., S. J. Bursian, R. H. Poppenga, et al. 1991. Toleration of high concentrations
of dietary zinc by mink. /. Vet Diagn Invest 3:232-237. As cited in Toxicological
Profile for Zinc, Agency for Toxic Substances and Disease Registry, U.S. Public Health
Service, Atlanta, GA, 1993.
Bleavins, M. R., R. J. Aulerich, J. R. Hochstein, et al. 1983. Effects of excessive dietary
zinc on the intrauterine and postnatal development of mink. Nutrition 113:2360-7. As
cited in Toxicological Profile for Zinc, Agency for Toxic Substances and Disease
Registry, U.S. Public Health Service, Atlanta, GA, 1993.
Davies, B. E., 1983. Heavy Metal Contamination from Base Metal Mining and Smelting:
Implications for Man and His Environment. Applied Environmental Geochemistry. ISBN
0-12-690640-8.
i
Davis, R.D. 1983. Crop Uptake of Metals (Cadmium, Lead, Mercury, Copper, Nickel, Zinc,
and Chromium) from Sludge-Treated Soil and its Implications for Soil Fertility and for the
Human Diet Proc. of the 3rd International Symp., Brighton, Sept.
Deuteh, B., B. Borg, L. Kloster, H. Meyer, and M. M. Moller. 1980. The accumulation of
£C_ ^
Zn by various marine organisms. Ophelia (Suppl 1):235-240.
;
Dewar, W. A., P. A. L. Wright, R. A. Pearson, and M. J. Gentle. 1983. Toxic effects of
high concentrations of zinc oxide in the diet of the chick and laying hen. Poultry Science
24:397-404.
Dowdy, R.H., W.E. Larson, J.M. Titrud, and J.J. Latterell. 1978. Growth and Metal Uptake of
Snap Beans Grown on sewage Sludge-Amended Soil: A four-Year Field Study. /.
Environ. Qual.,l(2) '• • '
August 1995
-------
APPENDIX B Zinc - 10
57 FR 24152. June 5; 1992. U.S. Environmental Protection Agency (FRL-4139-7). Draft
Report: A Cross-Species Scaling Factor for Carcinogen Risk Assessment Based on
Equivalence of mg/kg3/4/day.
Hale, J.G. 1977. Toxicity of metal mining wastes: Bull.Environ. Contain, and Tox. 17(1):66-
73.
Hare.L. 1992. Aquatic Insects and Trace Metals: Bioavailability, Bioaccumulation, and
Toxicity. Crit Rev. in Toxic. 22:327-369.
Helmke, P. A., W. P. Robarge, R. L; Kroter, and P. J. Schomberg. 1979. Effects of soil-
applied sewage sludge on concentrations of elements in earthworm. Journal of
Environmental Quality 8(3):322-327.
Hernmayer, K.L, P.E. Stake and R.L. Shippe. 1977. Evaluation of dietary zinc, cadmium, tin,
lead, bismuth and arsenic toxicity in hens. Poultry Sci. 56:1721-1722.
Honda, K., B.Y. Min, and R. Tatsukawa. 1986. Distribution of heavy metals and their age-
related changes in the Eastern Great White Egret, Egretta alba modesta, in Korea. Arch.
Environ. Contam. Toxicol:: 15, 185-197.
JamilJK. and S. Hussain. 1992. Biotransfer of Metals to the Insect Neochetina eichhornae via
Aquatic Plants. Arch. Environ.Contam. Toxicol. 22:459-463.
Khan.A.T. and J.S. Weis. 1993. Bioaccumulation of Heavy metals in Two Populations of
Mummichog (Fundulus heteroclitus). Bull. Environ. Contam. Toxicol. 51:1-5.
Kiffney, P.M. and W.H. Clements. 1993. Bioaccumulation of Heavy Metals by Benthic
Invertebrates at the Arkansas River, Colorado. Environ. Toxic, and Chem. 12:1507-1517.
Luckey, T.D. and B. Venugopal. Metal toxicity in mammals (1): Physiologic and chemical
basis for metal toxicity. Plenum Press, N.Y.
Nagy, K.A 1987. Field metabolic rate and food requirement scaling in mammals and birds.
Ecol.Mono. 57:111-128.
Opresko, D.M., B.E. Sample, G.W. Suter II. 1994. Toxicological Benchmarks for Wildlife:
1994 Revision. ES/ER/TM-86/R1. U.S Department of Energy, Oak Ridge National
Laboratory, Oak Ridge, Tennessee.
August 1995
-------
APPENDIX B Zinc - 11
Norberg, T. J., and D. I. Mount. 1985. A new fathead minnow (Pimephales promelas)
subchronic toxicity test. Environmental Toxicology and Chemistry 4(5):711-718. As
cited in AQUIRE (AQUatic Toxicity Information REtrieval Database). Environmental
Research Laboratory, Office of Research and Development, U.S. Environmental Protection
Agency, Duluth, MN.
Pal, N., and B. Pal. 1987. Zinc feeding and conception in the rats. Int J. Vitam. Nutr. Res.
57:437-440. As cited in Toxicological Profile for Zinc, Agency for Toxic Substances and
Disease Registry, U.S. Public Health Service, Atlanta, GA, 1993.
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Toxicity and production of abnormalities in development Ann. N.Y. Acad. Sci. 55:203.
Samanta, K., and B. Pal. 1986. Zinc feeding and fertility of male rats. International Journal
of Vitamin and Nutrition Research 56:105-107.
Schlicker, S.A. and D.H.Cox. 1968. Maternal dietary zinc, and development and zinc, iron,
copper content of the rat fetus. J Nutrition, 95: 287-294.
Smith, B. L., and P. P. Embling. 1984. The influence of chemical form of zinc on the
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Health and Environmental Assessment, Washington, DC.
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BulLEnviron.Contam.Toxicol. 38:580-587.
August 1995
-------
APPENDIX B Zinc - 12
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Duluth, MN.
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84OR21400. Office of Environmental Restoration and Waste Management, U.S.
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Biomagnification of Metals in a Pre-Cambrian Shield Lake. Water, Air and Soil Pollution
19:277-291.
August 1995
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Terrestrial Toxicity - Zinc
Cas No. 744O66-6 .
Chemical
Name
zinc
zinc oxide
zinc sullate
zinc sullate
zinc
carbonate
zinc sulfale
zinc (zinc
sultale)
Species ~
rats
rats
rats
minK
rat
mink
sheep
Endpoint
eP
systemic
rep
reD
"r
dvp
systemic
mortality
Description
PEL
NOAEL
PEL
NOAEL
LOAEL
NOAEL
PEL
Value
0.017
1,413
290
21.0
2.2
102.9
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
mg/kg-day
240 mq/kg-day
Exposure
Route (oral,
s.c.J.v., i.p..
injection)
oral
oral
oral
oral
oral
oral
oral
Exposure
Duration
/Timing
30-32 days
21 days
gestation
days 0-1 8
25 weeks
gestation
days 0-20
postpartum
days 70-2 14
3x/week for
four weeks
Reference
Samanta and Pal , 1986
-
Ansarjetal., 1976
Pal &Pa[, 1987
Bleavins et al., 1983 as
cited in ATSDR, 1993
yriu-Hareetal., 1989
Aulerich et al., 1991 as
cited in ATSDR, 1993
Smith and Embling, 19B4 as
cited in U.S. EPA, 1987
Comments
Reduced fertility of the males,
decreased sperm motility, and
reduced fertilising capacity were
all observed at this dose level.
Rats fed from 1 ,200 to 8,400
ppm zinc oxide did not exhibit
clinical symptoms of toxicity
such as skin lesions, diarrhea,
muscular incoordination, and
reduced feed intake.
Increased preimplantation
losses.
No effects on gestational length
or litter size.
There were effects on •
ossification centers, fetal length.
fetal weight, and number of
resorptions per liner at this dose
level.
No adverse pathological
changes in the kidneys, liver or
blood chemistry.
All animals died after day 13.
i
-------
Terrestrial Biological Uptake Measures - Zinc
Cos No. 7440-66-6
Chemical
.Name
zinc
zinc
zinc
NS = Not spa
Species
plant
earthworms
earthworms
cified
B-faclor
(BCF. BAF.
BMP)
BCF
BCF
BCF
Value
0.25
.68-5.4
2-3
Measured
or
Predicted
(m.p)
P
rn
NS
units
(ug/g OW
plant)/(ug/g soil)
NS
NS
Reference
US. EPA, 1990e
Oavies, 1983
Helmke. 1979
Comments
Data obtained from various
distances from point of soil.
-------
Freshwater Biological ->.e Measures - Zinc
CasNo./-.4U-66-6
Chemical
Name
zinc
zinc
zinc
zinc
zinc
NS = Not spe
Species
bluegill
3-spine
stickleback
15-spine
stickleback
fish
bass
cified
B-factor
(BCF, BAF,
BMP)
BAF
BCF
BCF
BCF
BAF
Value
245
130.00
200.00
47
803
Measured
or
Predicted
(m,p)
P
NS
NS
m
P.
Units
NS
NS
NS
-
Ukg
N.S
Reference
U.S. EPA, 1993
Deutch et al , 1980 as cited
in AQUIRE. 1994
Oeutch et at, 1980 as cited
in AQUIRE. 1994
U.SJiPA, 1992
Stephan, 1993
Comments
BAF calculated from data presented by
Murphy el al; 1978
Normalized to 3% lipid. As adjustment
was uncertain, could not use to derive a
whoje-body fish BCF.
BAF calculated from data presented by
Murphy etal.. 1978.
-------
M:
Freshwater Toxicity - Zinc
COS No. 7440-66-6
Chemical
Name
zinc
zinc
zinc
zinc
zinc
Species
aquatic
organisms
fathead
minnow
fish
daphnid
fish
NS = Not specified
Type of
Effect
chronic
acute
chronic
chronic
acute
Description
AWQC
LC50
CV
CV
EC20
Value
110
238 - 2540
(776)
36.41
46.73
47
Units
ug/L
ug/L '
ug/L
"9A
ug/L
Test Type
(Static/Flow
Through)
NS
NS
NS
NS
NS
Exposure
Duration
/Timing
NS
NS
NS
NS
NS
Reference
52 FR 62 13 403/02/87)
Norberg & Mount, 1985 and
Hobson and Birge as cited in
AQUIRE, 1995
Suteretal., 1992
Suteretal.. 1992
Suter et al.. 1992
Comments
-------
Terrestrial I .'/ - Zinc
Cas No. / -.40-66-6
Chemical
Name
.
zinc (zinc
oxide)
zinc (zinc
oxide)
zinc (zinc
oxide)
zinc (zinc
acetate)
oxide)
Species
rats
chickens
chickens
chickens
chickens
NS = Not specified
Endpoint
reprq
systemic
systemic
repro
pancreas
Description
NOAEL
NOAEL
LOAEL
NOAEL
LOAEL
Value
202
79 _
140
55.6
70
Units
mg/kg-day
mg/kg-day
mg/kg-day
mg/k^-day
mg/kg-day
Exposure
Route (oral,
s.c., i.v., i.p.,
injection)
oral
orcl
oral
oral
oral
Exposure
Duration
/Timing
36 days
42 days
42 days
56 days
42 days
Reference
Schlicker and Cox, 1968
Dewaretal , 1983
Dewar at al., 1983
*"
Hemmayer et al ,1952
Dewaretal.. 1983
Comments
No increase in the percentage
of fetal resorptions were seen at
this dose.
No effects on body weight or'
the gizzards of chickens were
observed at this dose level.
(1000ppm)
Reduced body weight and
increased gizzard erosion were
observed at this dose level.
(2000 ppm)
Could not use this study as a
benchmark because 1 ) data
were not included in published
abstract, 2) broken line
regression was used to derive
the NOAEL presented.
Asssumption made that weight
of 154 day hen equal to 140 day
hen (EPA, 1988),
exhibited pancreatic lesions.
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